&EPA
United States
Environmental Protection
Agency
Robert S. Kerr
Environmental Research Laboratory
Ada OK 74820
EPA 600/2-87/008
January 1987
Research and Development
Leaking Underground
Storage Tanks:
Remediation with
Emphasis on In Situ
Biorestoration
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EPA/600/2-87/008
January 1987
Leaking Underground Storage Tanks:
Remediation With Emphasis On In Situ Biorestoration
by
J. M. Thomas, M. D. Lee, P. B. Bedient, R. C. Borden,
L. W. Canter and C. H. Ward
National Center For Ground Water Research
Rice University, Houston, Texas 77251
University of Oklahoma, Norman, Oklahoma 73019
Oklahoma State University, Stillwater, Oklahoma 74078
Cooperative Agreement No. CR-812808
Project Officers
Marion R. Scalf and Jerry N. Jones
Applications and Assistance Branch
Robert S. Kerr Environmental Research Laboratory
Ada, Oklahoma 74820
U,3, Environmental Protection Agency
R"j:o!i 5, Library (PL-'PJ)
7/ West Jackson Boulevard, 12th Floor
Chicago, 1L 60604-3590
Robert S. Kerr Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Ada, Oklahoma 74820
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Disclaimer
The information in this document has been funded wholly or in part by the United States Environmental
Protection Agency under Assistance Agreement No. CR-812808 to Rice University and the National Center for
Ground Water Research. It has been subjected to the Agency's peer and administrative review, and it has been
approved for publication as an EPA document.
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Foreword
EPA is charged by Congress to protect the Nation's land, air and water systems. Under a mandate of national
environmental laws focused on air and water quality, solid waste management and the control of toxic substances,
pesticides, noise, and radiation, the Agency strives to formulate and implement actions which lead to a compatible
balance between human activities and the ability of natural systems to support and nurture life.
The Robert S. Kerr Environmental Research Laboratory is the Agency's center of expertise for investigation of
the soil and subsurface environment. Personnel at the laboratory are responsible for management of research
programs to: (1) determine the fate, transport and transformation rates of pollutants in the soil, the unsaturated
zone and the saturated zone of the subsurface environment; (2) define the processes to be used in characterizing
the soil and subsurface environment as a receptor of pollutants; (3) develop techniques for predicting the effect of
pollutants on ground water, soil and indigenous organisms; and (4) define and demonstrate the applicability and
limitations of using natural processes, indigenous to the soil and subsurface environment, for the protection of this
resource.
This project was initiated to provide a state-of-knowledge document of remediation measures applicable to
contamination resulting from leaking underground storage tanks. It has been estimated that more than three
million underground storage tanks exist and 10 percent or more are leaking contaminants into the subsurface. The
information found in this document includes sections on physical and hydrodynamic containment, withdrawal and
treatment methods, in situ physical, chemical and biological treatment, hydrologic considerations and
mathematical modeling of biorestoration, and institutional barriers to ground water pollution control. As the title
indicates, emphasis is placed on biorestoration with topics including: (1) microbial activity in aquifers, (2)
biostimulation by addition of limiting nutrients, (3) addition of specialized microbial populations to the
subsurface, and (4) enrichment of specific microbial populations. This compilation will be useful to those having
a need for information concerning possible remedial measures for ground waters contaminated by a variety of
sources.
Li/-
Clinton W. Hall
Director
Robert S. Kerr Environmental
Research Laboratory
111
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Abstract
Leaking underground storage tanks have been identified as a major source of ground water contamination in the
United States. More than three million underground storage tanks exist and 10 percent or more are leaking
compounds such as gasoline, process chemicals, hazardous chemicals, and dilute wastes into the subsurface. As a
result of the 1984 amendments to the Resource Conservation and Recovery Act (RCRA), regulation of
underground storage tanks now applies to tanks which have at least 10 percent of their volume underground and
which store petroleum products or hazardous liquid materials. Remedial techniques to restore contaminated
aquifers have been developed because as many as 300,000 cases of leaks have been estimated in the United States.
The state-of-the-art knowledge of measures available for remediation of contaminated ground water that results
from such leaking underground storage tanks has been provided in this document. In situ biorestoration
techniques are emphasized.
Treatment of contaminated aquifers has traditionally included physical containment, hydrodynamic control, and/
or withdrawal and treatment. In situ physical and chemical treatment schemes have also been developed.
The current literature indicates that in situ biorestoration has great potential for remediation of contaminated
aquifers. In situ aquifer restoration involves the enhancement of the indigenous microflora to degrade subsurface
pollutants. A second but unproven stategy is to add specialized bacteria to the subsurface. Aquifers amenable to
biorestoration are usually those that are perfusable with solutions which carry the nutrients to the zone of
contamination. The presence of indigenous microorganisms that can degrade subsurface contaminants has been
demonstrated. However, a period of adaptation is usually required before the subsurface microflora can degrade
the pollutants. Factors which may limit degradation, even in the presence of adapted organisms, include lack of
an essential nutrient, substrate concentration, substrate inacces-
sibility, pH, temperature, and the presence of toxicants. Biorestoration of contaminated aquifers often involves
the addition of limiting nutrients such as oxygen, nitrogen, and phosphorus. Enriching for microorganisms with
special metabolic capabilities has been demonstated in the laboratory and may be promising in in situ
biorestoration schemes. Mathematical models of biorestoration have been developed to simulate progress of the
cleanup and provide information on the kinetics of the process.
IV
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Contents
Disclaimer.. ;.....„....„ ii
Foreword ....„ iii
Abstract iv
List of Tables and Figures vii
I. Introduction 1
A. Importance of Ground Water Protection 1
B. Universal Impact of Leaking Underground Storage Tanks 1
C. Definition-Underground Storage Tanks (RCRA) 2
D. Available Remedial Action Technology 3
E. Subsurface Effects on Contaminant Mobility 3
II. Remedial/Restoration Plume Management Techniques 4
A. Physical Containment 4
1. Removal 4
2. Barriers to Ground Water Flow 4
3. Surface Water Controls 8
B. Hydrodynamic Controls 9
1. Passive Ground Water Controls 9
2. Well Systems 10
C. Withdrawal and Treatment 11
1. Chemical Treatment 11
2. Physical Treatment 14
3. Biological Treatment 19
D. In Situ Physical and Chemical Treatment 22
E. In Sku Biological Treatment 24
1. Microbial Activity in Aquifers 24
2. Biostimulation by Addition of Limiting Nutrients 29
Development of the Jn Situ Biostimulation Process With Oxygen
Supplied by Air Sparging 29
Alternative Oxygen Sources 37
Summary of Aerobic In Situ Biostimulation Processes 42
Innovative Processes 45
Potential for Anaerobic Processes 49
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3. Addition of Specialized Microbial Populations to the Subsurface 51
4. Enrichment of Specific Populations 58
F. Hydrologic Consideration and Mathematical Modeling of Biorestoration 62
1. Hydrologic Considerations 62
2. Modeling Biorestoration 64
3. Kinetics of Biodegradation 64
4. Modeling of Subsurface Transport 67
5. Mathematical Models of Subsurface Biorestoration 69
6. Model Use and Limitations 71
7. Conclusions 72
III. Institutional Limitations on Ground Water Pollution Control 73
A. Scientific Understanding of the Nature of Released
Products from Leaking Underground Storage Tanks 73
1. Physical and Chemical Nature of Petroleum Products 73
2. Nature of Occurrence of Leaks 74
3. Hydrologic Complications 75
B. Public Opinion 76
C. Business Community Attitudes 77
1. UST Owners and Operators 77
2. Insurance Companies 78
3. Petroleum Product Sales Support and Maintenance Companies 78
4. General Business Community 79
D. Environmental Interest Groups 79
E. Government Agencies 79
1. Unclear Jurisdictions 79
2. Who is Responsible for Clean Up? 80
3. What is an Adequate UST Program? 80
4. How Clean is Clean? 80
5. General Institutional Concerns 81
IV. Research Needs for Optimized Remedial Techniques 81
A. Evaluation of Effectiveness of Physical Containment Techniques 81
B. Enhanced Vadose Zone Pollutant Removal Techniques 82
C. Enhancement of Microbial Populations 82
V. References 83
VI
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List of Tables
Number Page
2-1 Advantages and Disadvantages of Some Physical Barriers to Ground Water Flow
(Hill et al., 1980; Carter and Knox, 1985; Brunsing and Cleary, 1983;
Nielsen, 1983; Truett et al., 1982) 5
2-2 General Characteristics of Some Chemical Treatment Processes (Law Engineering, 1982) 12
2-3 Organic Compounds That Have Been Shown to be Biodegradable in the Subsurface 27
2-4 Advantages and Disadvantages of Biorestoration (J. R. B. Associates, 1982; Yang and Bye, 1979) 42
2-5 Contaminants Treated by In Situ Biostimulation 43
2-6 Types of Aquifers Where In Situ Biostiumulation Has Been Utilized 46
2-7 Factors That Control Biodegradation in Land Treatment (Huddleston, et al., 1986) 47
2-8 Reasons Why Introduced Organisms Fail to Function in the Environment (Goldstein et al., 1985) 53
2-9 Summary of Aquifer Remediation Case Histories Utilizing Introduced Organisms 56
2-10 Estimated Quantities of Oxygen and Methane or Propane Required to Bring the Concentration of
Trichloroethylene, cis- or trans-1,2-Dichloroethylene or Vinyl Chloride to 5 )J,g/L 60
2-11 Prospects for Treatment of the Common Halogenated Organic Contaminants in Aquifers Through
Co-oxidation Supported on Gaseous Alkanes 61
List of Figures
2-1 Typical schematic for aerobic subsurface biorestoration 33
2-2 Use of infiltration gallery for recirculation of water and nutrients in in situ biorestoration 33
2-3 Combination of above ground treatment with in situ biorestoration 57
vn
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I. Introduction
A. Importance of Ground Water Protection
Ground water is an important resource in the United States due to its current and expected increase in use as a
water supply for domestic, industrial, and agricultural purposes. Two major issues of long-range concern include
ground water depletion from over usage, and ground water quality impairment due to man-made pollution
sources. Management of withdrawals to coincide with natural rates of recharge can minimize depletion concerns.
Existing quality impairment can be alleviated by remediation techniques and preventative measures can be used to
preclude further declines in the quality of the resource. Measures taken to prevent pollution of ground water are
the best techniques for aquifer preservation.
One ground water pollution source category which has recently received increased attention is underground
storage tanks, specifically, the inadvertent leakage of liquid products from such tanks into the unsaturated and
saturated zones. This document was requested by the U. S. Environmental Protection Agency (EPA) Office of
Underground Storage Tanks (OUST) to aid in planning a program for restoration of ground water contaminated by
leaking underground storage tanks. The report is intended to be a state-of-the-art document to guide OUST in
making judgements regarding the efficacy of alternative technologies for remedial action. The report specifically
focuses on the use of innovative biological technologies to remove or attenuate contaminants.
B. Universal Impact of Leaking Underground Storage Tanks
One of thjejnmary_soarces_oi[ jromidjwatgr contamination has bemidmtififid-asJeaking underground tanks which
ej fuels, pjo£ejs_cliemicals, hazaidpjus^?AJoxic_cJiejTiicah.and dilute wastes^
_
(Cherimisinoff, Casana, and Ouellette, 1986). Snow (1985) has noted that of all the sources of ground water
contamination, leaking underground storage tanks may pose the most serious threat to human health. The total
number of underground storage tanks in the United States is probably upwards of three million. A conservative
estimate is that 10 percent or more of these tanks are leaking, thus as many as 300,000 incidents of ground water
contamination may be occurring (Dowd, 1984). Feliciano (1984) indicated that there are about 1.4 million
underground tanks storing gasoline, with 85 percent made of steel and having no corrosion protection. The
majority were buried more than 20 years ago. Some petroleum experts estimate that 75,000 to 100,000 of these
tanks may be leaking their contents into ground water (Feliciano, 1984).
A national survey of underground motor fuel storage tanks has been conducted at 890 establishments covering
2,445 tanks (Klein, 1986). A subsample of 433 tanks was tested for leaks, with the following pertinent findings:
(1) An estimated 35 percent of the tank systems failed the tightness test. This conclusion must be interpreted with
caution, however. In EPA's test, tank systems were filled above normal operating levels, which slightly
elevated the normal operating pressure on the tank.
(2) The average rate for leaking tanks, adjusted for typical operating conditions, was 0.31 gallons per hour. Half
the leaks were 0.25 gallons per hour or less. In the statistical analysis of this survey, EPA could not identify
any variable (such as type of material or fuel type) that strongly correlated with test failure.
(3) At comparable ages, fiberglass and steel tanks showed no significant difference in the rate of test failure.
Steel tanks comprise an estimated 89 percent of all underground motor fuel storage tanks, with fiberglass
making up the rest. Steel tanks showed little increase in the rate of test failure with age in their first 20 years,
but after 20 years failures increased.
(4) There are an estimated 796,000 individual motor fuel storage tanks, located at an estimated 326,000
establishments.
(5) The average age of tanks in the United States is 12 years.
(6) Twenty-one percent of the tanks are installed partially or completely below the water table.
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In summary, underground storage tanks are ubiquitous and many are leaking. Existing tanks which are not
currently leaking are expected to do so as they age. Therefore, there is a need for regulatory programs to protect
ground water from this pollutant source.
C. Definition-Underground Storage Tanks (RCRA)
The Hazardous and Solid Waste Amendments to the Resource Conservation and Recovery Act (RCRA) of 1984
(P.L. 98-616) added a new Subtitle I, "Regulation of Underground Storage Tanks". Part of these amendments
require the EPA to develop and establish a national regulatory program for the control of new and existing
underground storage tanks (U.S. EPA, 1985). The scope of this new program is broad and applies to tanks and
combinations of tanks with 10 percent or more of their volume underground, including the volume of underground
piping, that are used to store petroleum products or other liquid materials defined as hazardous substances under
Section 101(14) of the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA,
commonly known as Superfund). The following tanks are excluded from the Interim Prohibition:
(1) farm or residential tanks having a capacity of 1,000 gallons or less for storing motor fuel for noncommercial
purposes;
(2) tanks used for storing heating oil for consumptive use on the premises where stored;
(3) septic tanks;
(4) flow-through process tanks; and
(5) tanks above floor level but still underground.
Underground tank systems may leak due to several reasons. Corrosion, both external and internal, is considered
the most common cause of leaks. Structural failure, primarily from improper tank installation, can also cause
leaks. In addition, tank contents that are incompatible with a tank's liner and/or construction materials may induce
leakage (Plehn, 1984). Poor operating practices have also been noted as a cause of tank leakage (Woods and
Webster, 1984).
Existing tanks are to be addressed by an inventory and monitoring program. For example, among Subtitle I's
provisions is the requirement that by May 8, 1986, underground tank owners notify designated state or local
agencies of their tanks' existence. Under the law, EPA must also develop regulations for underground tanks
addressing leak detection, corrective action, closure, record keeping and reporting, and new tank performance
standards (Miller and Taylor, 1985). In addition, any new tank installed after May 7, 1985, must be designed,
installed and operated so that it:
(1) will prevent releases due to corrosion or structural failure for the operational life of the tank;
(2) is cathodically protected against corrosion, constructed of noncorrosive material, steel clad with a
noncorrosive material, or designed in a manner to prevent the release or threatened release of any stored
substance; and
(3) the material used in the construction or lining of the tank is compatible with the substance to be stored.
In its interpretive rule, EPA has noted that the term "operational life" of a tank is "the time during which the
tank stores regulated substances." The Interim Prohibition also has a limited exemption stating that a new tank
does not have to be protected from corrosion if it is installed in a certain soil environment as follows:
j^ or anotner standard approved
"By the AdrnTnMrator7show that soil resigjivijyjn an installation location is 12,000 ohm/cmjor
rnorejunless a more stringent standard is prescribed by the Administrator by rule), a storage tank
without corrosion protection may be installed in that location.
Approximately one-half of the 50 states have initiated regulatory programs for underground storage tanks. In
addition, local governmental agencies have developed regulatory programs (Zakheim and Ehrhardt, 1985).
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D. Available Remedial Action Technology
Remedial action technologies for leaking underground storage tanks can be considered in relation to preventive
measures and contaminant plume management. Preventive measures include corrosion prevention, proper
installation practices, material, selection, and containment systems. Detailed information concerning preventive
measures can be found elsewhere (National Technical Information Service, 1985; National Fire Protection
Association, 1984; American Petroleum Institute, 1979; U. S. EPA, 1985). Plume management techniques may
include one or several of the following (Repa and Doerr, 1986):
(1) ground water pumping to extract water from or inject water into wells to capture a plume or alter the direction
of ground water movement;
(2) subsurface drains consisting of permeable barriers designed to intercept ground water systems;
(3) vertical underground barriers made of low-permeability materials to divert ground water flow or minimize
leachate generation and plume movement; and
(4) innovative technologies that biologically or chemically remove or attenuate contaminants in the subsurface.
The most recent and innovative techniques for aquifer remediation involve stimulating the indigenous microflora
to remove subsurface pollutants. Stimulation of the native microorganisms can result in the complete destruction
of the contaminants whereas chemical or physical treatment may result in incomplete destruction or transfer of the
contaminant to another phase of the environment. Biostimulation is often achieved by adding limiting nutrients
such as oxygen, nitrogen, and phosphorus. The addition of specialized microbial populations to degrade
subsurface pollutants has been incorporated into many remedial projects; however, their role in biodegradation
could not be distinguished from that of the indigenous microflora.
E. Subsurface Effects on Contaminant Mobility
The transport and fate of products released to the^subsurface environrnent from leaking underground storage tanks
are dependent upon hydrpdynamic. abiotuLand biotic processes. A basic understanding of the nature and
influences of sthese processes on contaminant mobility is necessary in the identification and implementation of
appropriate remedial action technologies. The relative importance of the three basic processes is dependent upon
both product (contaminant) characteristics and local hydrogeological features.
Some studies have been done on these processes as related to leaking underground storage tanks, and others are in
progress. Bjotojrreessj^ obviously important for certain organic products,
including many components of petroleum. For example, Jenson (1985) found positive evidence for natural
biodegradation of gasoline which had leaked from an underground storage tank. In another example, ground water
contamination was detected 30 m beneath a paint factory in Milan, Italy (Botta, Castellani, and Mantica, 1983).
The pollution resulted from the leakage of organic solvents from underground storage tanks. Extraction of water
samples at various pH values yielded three fractious: neutral, acidic, and basic. Aliphalic and aromatic acids were
found in the acidic fraction, although their presence could not be attributed to industrial use. Observed
concentrations of these and other compounds were attributed to oxidative microbial degradation of hydrocarbon
components present in soil and ground water.
H^d^lyranii£2ro£essg&,can also be involved in contaminant mobility. For example, the rate of movement of
leaked oil into soil will d^gjd^njrodu£tjvisc^ty_and_soil properties. Light products, such as gasoline, will
penetrate rapidly, while heavy oils will move more slowly (American Petroleum Institute, 1980). The rate of
penetration is also a function of the soil permeability. In clays with low permeability the product may penetrate
very little, if at all; conversely, in sandy soil the penetration may be rapid. In the unsaturated zone, some oil may
become trapped between individual soil particles, remain behind the main body of oil, and provide slow leakage
of contaminants over a long period of time. This trapped or residual oil can remain in place for long periods of
time. Residual hydrocarbons can occupy from 15 to 40 percent or more of the pore space as a result of these
trapping processes depending on several physical characteristics of the subsurface. The ability to design and
conduct successful remediation strategies depends largely on the ability to understand, predict and enhance the
mobility of both liquid and dissolved hydrocarbons (Wilson and Conrad, 1984). Theoretically, there are a number
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of ways the trapped residuals can be mobilized and concentrated such that a much lower content of residual
product will have to be dealt with. Obviously, the most used mobilization technique is to increase the hydraulic
gradient, usually by pumping, thereby increasing the Darcy velocity of the water phase in the saturated zone.
Light products can move to the surface of the water table in unconfined aquifers and spread out along this surface,
occupying some of the capillary zone as well as the top layers in the water table (American Petroleum Institute,
1980). Finally, many oils and refined products contain components which are slightly soluble in water and other
components which might volatilize. Solubility is greatest with lighter, aromatic components; these components
are then influenced by advection and dispersion processes in the saturated zone. Soluble components can include
benzene, toluene, and xylene. The mathematical modeling of multicomponent organic mixtures from leaking
underground storage tanks is beyond the scope of this report; however, some work has been done in this area
(Uchrin and Slater, 1985).
_AWoticjproce^ses can also influence contaminant mobility in the subsurface. These processes include adsorption,
ion^exchange, chemicalprecipitation and jomplejajtion^Adsorpdon of organics increases product retentionTwith
the adsorption being a function of the octanol-water partition coefficient of the chemical and the inherent organic
content of the subsurface media. Transport of metals from leaking underground storage tanks may be influenced
by one or more of the above processes. The process that dominates depends upon the metal, soil characteristics,
and subsurface environmental conditions such as pH and the oxidation-reduction potential.
II. Remedial/Restoration Plume Management Techniques
A. Physical Containment
Physical containment measures isolate contaminated soil and ground water from the local environment and
minimize any threat to human health (Wright and Caretsky, 1981). Isolation techniques for the surface and
subsurface include excavation and removal of the contaminated soil and ground water, barriers to ground water
flow, and surface water controls. In a survey of 169 remedial actions by Neely et al. (1981), in situ containment
and removal were the most frequently used options.
1. Removal
Excavation and removal of contaminated soil and ground water may be used when in situ containment or
treatment is unacceptable (Wright and Coretsky, 1981). A pit is dug to remove the soil or pumping wells are
installed to control the plume (Josephson, 1980). The excavated soil is transported to a secured site, such as a
landfill or surface impoundment, for disposal (Ehrenfeld and Bass, 1984). The ground water is pumped out and
can be treated using a variety of techniques. The cost of excavation of soil ranges from $1.75 to $4.50 per cubic
yard. The inherent problem in excavation and removal to another location is that a new hazardous waste site is
often created. Several Resource Conservation and Recovery Act (RCRA) facilities that have received
contaminated materials from Superfund cleanups are leaking and are now new Superfund sites (Hileman, 1984).
However, removal of contaminated soil and ground water to a more environmentally appropriate location may be
necessary if in situ, containment or treatment poses health risks or initiates litigation. Another problem associated
with excavation and removal is that total removal of subsurface soil and ground water may be impossible when
the contamination extends deep beneath the surface or the contaminants lay below an immobile facility.
2. Barriers to ground water flow
Physical barriers used to prevent the flow of ground water include slurry walls, grout curtains, sheet piling, block
displacement and clay liners (Ehrenfeld and Bass, 1984). These impermeable barriers may contain contaminated
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contaminated ground water or leachate or prevent the flow of clean ground water into a zone of contamination.
Some general characteristics of slurry walls, grout curtains, sheet pilings and block displacement are compared in
Table 2-1.
Table 2-1. Advantages and Disadvantages of Some Physical Barriers to Ground Water Flow (Hill et al., 1980;
Canter and Knox, 1985; Brunsing and Cleary, 1983; Nielsen, 1983; Truett et al., 1982)
Methods
Advantages
Disadvantages
Slurry walls
Grout curtains
very low upkeep
fairly effective
simple construction
long service life
minimal environmental
impact
very low upkeep
fairly effective
increases soil's bearing
capacity
long service life
minimal environmental
impact
used in consolidated and
unconsolidated material
versatile
specific targets can be
reached
a variety of fluids can
be injected
expensive
bentonite deteriorates
in concentrated ionic
solutions
must be anchored to
impermeable strata
some construction
procedures patented
bentonite availability
limited
very expensive
hard to place
difficult to determine
completeness of wall
some techniques are
proprietary
limited to soils with
permeabilities of
10~5 cm/sec or
greater
cannot be used at shallow
depths (1.5 m)
some applications can
create additional
pollution
Sheet pilings
inexpensive for small
projects
fairly effective
minimal environmental
impact
long service life
sections of steel piling
are reusable
difficult to form
effective barrier
in coarse, dense
material
steel may corrode from
contamination
not initially watertight
(Continued)
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Table 2-1. (Continued)
Methods Advantages Disadvantages
Block displacement economical when impermeable in developmental stages
layer is deeper than zone
of contamination
isolates minimum amount of
soil
bottom barrier is formed
without excavation
Slurry Trench Cutoff Walls--
Aquifers with sandy surficial soil less than 60 ft in depth and underlain by an impermeable layer of fine grain
deposits or bedrock are most amenable to slurry wall construction (Need and Costello, 1984). Construction of a
slurry wall entails excavating a narrow trench (2 to 5 ft) surrounding the contaminated zone. The slurry acts to
maintain the trench during excavation and is usually a mixture of soil or cement, bentonite, and water (Ehrenfeld
and Bass, 1984). The trench is generally excavated through the aquifer and into the bedrock. Installation of a
slurry wall at depths greater than 60 ft is difficult.
There are two different methods for construction of slurry walls (Tallard, 1984). Trenches constructed using a
cement-bentonite (C-B) mixture are allowed to set whereas those constructed with a soil-bentonite (S-B) mixture
are backfilled and solidified with appropriate materials. Solidification of the trench may be accomplished by
backfilling with soil mixed with bentonite, soil mixed with cement, concrete, an asphaltic emulsion, or a
combination of these with synthetic membranes (Lynch et al., 1984). The chosen materials should be compatible
with the in situ soil and contaminant regime. Depending on the backfill material used, the permeability of the
resulting barrier may range from 10"6 to 10"8 cm/sec (Nielson, 1983). A slurry wall isolation system may be
constructed by placing a low permeability cap over a waste site that has been completely surrounded by a slurry
wall (Need and Costello, 1984). Soil-bentonite trenches are commonly used in the United States (Tallard, 1984;
Lynch et al., 1984). Further details on the construction, permeability, compressibility, and strength of a S-B wall
in addition to the effects of pollutants on wall permeability are given in a paper by D'Appolonia (1980). Both S-B
and C-B trenches are stable under normal conditions; however, C-B trenches deteriorate in the presence of acids
and sulfates and S-B trenches are similarly affected by organics, calcium, magnesium, heavy metals, and
concentrated ionic solutions (Canter and Knox, 1985). Laboratory tests should be conducted to determine the
appropriate materials that should be used under the existing contaminant regime (Lynch et al., 1984; Tallard,
1984).
A S-B slurry cut-off wall was chosen for the first clean up financed by the Environmental Protection Agency's
Superfund program (Ayres et al., 1983). The Gilson Road uncontrolled hazardous waste disposal site was an
abandoned sand and gravel pit in which drums, chemical sludges, demolition debris, and municipal rubbish were
illegally dumped. The underlying aquifer was found to contain a contaminant plume more than 1,500 feet in
length, 110 ft in depth, with an areal extent of 30 acres. After removal of the drums, the site was filled, graded,
and a S-B slurry wall isolation system installed. The design of the S-B backfill material is discussed in detail by
Schulze et al. (1984). The cap was constructed with 40-mil high-density polyethylene (Ayres et al., 1983).
The use of synthetic liners in slurry trenches is an old concept but largely undemonstrated (Tallard, 1984). The
problem with lining slurry walls is the difficulty in installation of the liner; the liner is usually welded at the
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surface and placed in the trench with complex rigging. However, a liner can be custom designed to resist selected
contaminants. One type of slurry wall liner system is constructed by backfilling a double liner sheet envelope
with porous fill; observation wells are placed in the fill material to monitor seepage. The Envirowall vertical
cutoff barrier is an example of such a hybrid cutoff wall (Arlotta et al., 1983). The trench is lined with high-
density polyethylene (HDPE) and backfilled with sand; the permeability of the HOPE envelope is 10'12 cm/sec.
Slurry walls may also be constructed using the vibrating beam method (Canter and Knox, 1985). The vibrating
beam wall is constructed by first excavating an 18-20 inch deep reservoir trench and then filling the trench with
slurry (Schmednecht and Goldbach, 1984). An I-beam attached to a pressure hose is vibrated through the trench
and into the ground. As the beam is withdrawn, slurry is injected into the ground under pressure. A vibratory
driver-extractor became commercially available in the late 1950s that both drives and extracts the beam
(Schmednecht and Goldbach, 1984). Slurries consisting of cement-bentonite or asphaltic emulsions may be used.
Asphaltic emulsion cut off walls are more expensive but less permeable (IxlO"8 cm/sec) and resistent to corrosive
chemicals than C-B slurries. In comparison to S-B slurry walls, the vibrating beam wall may be more cost-
effective because a smaller amount of grout is required to form the thin barrier (Tallard, 1984). However,
heterogeneities in the subsurface soil or the vibrations from the injection process may result in a discontinuous
wall; imperfections in the wall cannot be detected during the construction process.
Liners--
Liners, another type of physical barrier, are often used in conjunction with surface water controls and caps (Canter
and Knox, 1985). In addition to use in slurry walls, liners may be used to protect ground water from leachate
resulting from landfills containing hazardous materials. The type of liner used depends on the type of soil and
contaminants which are present. The average cost of liner installation per acre is $40,000 to $50,000.
Liners include polyethylene, polyvinyl chloride (PVC), many asphalt-based materials, and soil-bentonite or
cement mixtures (Canter and Knox, 1985). Polyvinyl chloride liners have permeabilities of less than 3.2 x 10'11
ft/sec; however, little is known about the service life of the PVC membranes (Threlfall and Dowiak, 1980). The
membrane should be installed in fine-grained soil to prevent punctures.
Grout Curtains--
Grout curtains are another type of physical barrier which are constructed by injecting grout (liquid, slurry, or
emulsion) under pressure into the ground through well points (Canter and Knox, 1985). Ground water flow is
impeded by the grout that solidifies in the interstitial pore space. The curtain is made contiguous by injecting the
grout into staggered well points that form a 2 or 3 row grid pattern (Ehrenfeld and Bass, 1984). Spacing of the
well points for grout injection depends on the radial extent of grout penetration. In addition, the rate of injection
is critical. Premature solidification occurs when the injection rate is too slow whereas the soil formation is
fractured when the rate is too fast. Soil permeability is decreased and soil-bearing capacity is increased after the
grout properly solidifies (Canter and Knox, 1985). Grouting is not effective in soils if 20 percent or more of the
particles are less than 20 mesh in size. The average cost of grouting ranges from $142 to $357/ft3.
Two types of grouting materials can be used: chemical and particulate grouts. Chemical grouts, consisting of two
or more materials that gel upon contact (Canter and Knox, 1985) may be constructed with sodium silicate,
phenoplasts, lignosulfonate derivitives, aminoplasts, or acrylamide (Ehrenfeld and Bass, 1984). The sodium
silicate grout is the most widely used chemical grout, can be used at slow injection rates in fine sediments, is
resistant to freezing and thawing, and reduces permeabilities from 10'2 to 10'8 cm/sec in sands. Phenoplasts and
lignosulfates are rarely used because of problems with toxicity and high cost. Chemical grouts are not appropriate
for acidic or basic soils and/or contaminants because the gelling process is an acid-base reaction.
Particulate, or suspension grouts, are a mixture of materials such as cement, bentonite, rock flour, clay, asphalt, or
sand with water and are best suited for coarse or gravel soils (Canter and Knox, 1985). Suspension grouts made
with Portland cement are widely used because they set within two hours (Nielsen, 1983). The composition of the
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grout which is selected depends on the type of soil contaminants present. Of all the grouting techniques,
particulate grouts are least expensive and most widely used.
Sheet Piling--
Construction of steel piling involves driving interlocking sections of steel sheet piling into the ground (Canter and
Knox, 1985). The sheets are assembled before use by slotted or ball-and-socket type connections and are driven
into the soil in sections. The piles are driven through the unsaturated zone, the aquifer, and down into the
consolidated zone using a pile hammer (Nielsen, 1983). After driving the barrier into the consolidated material,
the piles remaining above ground are usually cut off (Canter and Knox, 1985). The connections between the steel
sheets are not initially water tight; however, fine grained soil particles eventually fill the gaps and the barrier
generally becomes impermeable to ground water flow. Sheet piling may be ineffective in coarse, dense material
because the interlocking web may be disrupted during construction (Nielsen, 1983).
Sheet piling is usually effective to depths of 49 ft; however, deeper barriers have been reported (Ehrenfeld and
Bass, 1984). In addition to steel sheet piles, concrete also has been used. Concrete is expensive but is used when
a barrier stronger than the steel sheet piles is necessary. Steel sheet piles are limited in that they may be
susceptible to corrosion by some contaminants. The longevity of steel piling may be 40 years when the pH ranges
from 5.8 to 7.8; however, a pH of 2.3 may limit the barrier to seven years of service. Construction of a 1700 ft-
long by 60 ft-deep sheet pile was estimated to cost between $650,000 to $956,000 (Canter and Knox, 1985).
Block Displacement--
Block displacement is a method by which large areas of soil can be isolated by constructing an impermeable
barrier around and underneath a block of earth (Cleary, 1980). The technique is currently under development for
application to hazardous waste sites (Ehrenfeld and Bass, 1983). The bottom and surrounding barriers are
constructed in two separate processes (Brunsing and Cleary, 1983). The bottom barrier is constructed by pumping
a bentonite slurry into a series of injection wells. The subsurface material fractures as the slurry material from
different wells coalesces. Continued pumping under pressure vertically displaces the block of earth. The amount
of displacement corresponds to the thickness of the barrier. Construction of the surrounding barrier may be
accomplished using traditional techniques such as slurry walls and grouting (Brunsing and Cleary, 1983). The
surrounding barrier may be constructed first to induce a favorable stress field for formation of the bottom barrier
(Ehrenfeld and Bass, 1984). The two barriers must intersect, regardless of which is constructed first. The type of
slurry used for barrier construction depends on the same parameters considered for other types of physical
barriers, i.e., soil type, ground water, and contaminant regime (Ehrenfeld and Bass, 1984). Block displacement
may be preferable to slurry walls when the depth to the unconsolidated material is greater than 60 ft.
3. Surface Water Controls
Caps, dikes, terraces, channels, chutes, downpipes, grading, vegetation, seepage basins, and ditches are used to
divert uncontaminated surface water away from waste sites and reduce the amount of infiltration and resulting
hazardous leachate, or direct contaminated water away from clean areas (Ehrenfeld and Bass, 1984). Many of
these techniques may be used in conjunction with each other. Surface capping usually involves covering the
contaminated area with an impermeable material, regrading to minimize infiltration of surface water, and
revegetation of the site (Canter and Knox, 1985). For example, the remedial action taken at an abandoned
municipal/industrial dump that contained polychlorinated biphenyls (PCBs) included capping, revegetation,
surface and subsurface drainage, and a gas control system (Blasland et al., 1982).
Surface caps are usually constructed using materials from one of three groups: 1) natural soils, 2) commercially-
designed materials, or 3) waste materials (Canter and Knox, 1985). The material used should be compatible with
the soil type and contaminant regime. Examples include clay, concrete, asphalt, lime, fly ash, and mixed layers of
synthetic liners (Erdogan, 1984). Fine-textured soils, often from on-site, are most commonly used. The soil is
compacted into a cap that covers the waste site and minimizes infiltration of surface water (Farb, 1978). Blending
the soil with additives and cements may increase the effectiveness of the resulting cap (Canter and Knox, 1985).
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Using a combination of different materials, such as loam, clay, and gravel, in layers may also improve surface
water control. Commercially available materials such as bituminous or Portland cement concrete barriers and
various types of membranes may be more desirable than soil when the waste is extremely hazardous. The average
cost of cap installation ranges from $20,000 to $30,000/acre.
Dikes are used to control runoff or flood water whereas terraces are designed to divert surface water away from an
area or control erosion (Ehrenfeld and Bass, 1984). Downdrain structures, such as chutes and pipes, are used to
channel surface water downgradient (Doyle, 1980). Channels, or diversion ditches, are used to intercept runoff
upstream of a location and direct the flow along a different course (Doyle, 1980). Surface water intercepted by
channels may be diverted into seepage basins and ditches which discharge the water into the ground (Ehrenfeld
and Bass, 1984). Grading changes the slope and runoff characteristics of a site; soil compaction is the most
important method of grading. Revegetation may be appropriate after grading of a site to minimize erosion. Plants
may also take up hazardous materials from contaminated soils in addition to aesthetically improving the
appearance of the site. However, assimilation of hazardous materials by plants may pose environmental problems
by providing an entry of the contaminants into the food chain.
There are no special limitations in using surface controls; however, surface controls only isolate the contamination
rather than provide active treatment or management of plumes (Ehrenfeld and Bass, 1984). The availability of
appropriate materials for surface caps may limit this approach. The use of other surface water controls depends on
factors such as site topography, contaminant characteristics, and soil type.
B. Hydrodynamic Controls
Manipulation of the hydraulic gradient is one of the most commonly used methods for plume management and
remedial action at hazardous waste sites. Shepard (1983) reported that hydrodynamic controls were required in
most instances in which physical barriers such as slurry walls, grout curtains, and sheet piles were successful in
plume management. Further analysis indicated that the hydrodynamic control would have been just as successful
without the physical barrier.
The type of hydrodynamic control used may be passive or active. Passive hydrodynamic controls, or interceptor
systems, function by gravity flow, whereas pumping is required for active control (Canter and Knox, 1985).
1. Passive Hydrodynamic Controls
An interceptor system is constructed by digging a trench below the water table; the process creates a zone of
depression along the length of the trench (Canter and Knox, 1985). A perforated pipe is usually placed in the
trench which is then backfilled with coarse material to aid in the collection and transport of water. There are two
applications of interceptor systems, collector drains and interceptor trenches, which can be used as preventive or
abatement measures. Collector drains such as leachate collection systems, are preventive measures against ground
water contamination, whereas relief and interceptor drains are used to abate existing contamination. Relief drains
are used in formations with relatively flat hydraulic gradients to lower the level of the water table or prevent
contaminants from moving into deeper aquifers. Interceptor drains may be used to divert water away from a
waste site or prevent contaminated water from entering an uncontaminated area. The second type of interceptor
system, interceptor trenches, may operate by gravity flow or by pumping. The gravity flow, or passive interceptor
trenches, are mainly used to remove pollutants such as hydrocarbons that float on top of the water table. The
trenches remain open so that the contaminants can be removed with a skimming pump. The contaminants must be
removed continuously to prevent seepage into the trench wall and transport downgradient in the formation.
The obvious advantage to using passive hydrodynamic control is the low cost of operation (Canter and Knox,
1985). Construction is simple, flexible, and the resulting system is reliable. However, passive control is not
amenable to low permeable soils and deep aquifers, and usually cannot be installed underneath existing pits or
other facilities. In addition, passive interceptor systems require frequent monitoring and maintenance.
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2. Well Systems
Well systems are active measures of hydrodynamic control that rely on pumps or other mechanical devices for
plume management. Active hydrodynamic control is the most frequently used technique for aquifer restoration
(Jackson, 1982). The hydraulic gradient is controlled using withdrawal and/or injection wells (Canter and Knox,
1985). There are two types of withdrawal methods, well point and deep well systems. Well point systems are
used in shallow aquifers and are comprised of numerous closely spaced wells mat are centrally pumped. Deep
well systems are designed for deeper aquifers and differ from well point systems in that the deep wells are
pumped individually. Pressure ridge systems involve injection of water to manage plumes. Injection of water into
the aquifer causes an upconing of the original water table which prevents the flow of ground water in that area.
Design of any well system must be preceded by a thorough hydrogeological characterization of the site (Canter
and Knox, 1985; Nielsen, 1983). The number and placement of wells depends on the plume dimensions,
hydraulic gradient, and the hydrogeological characteristics of the aquifer. Plume management may be
accomplished using both withdrawal and injection wells (Nielsen, 1983). A row of withdrawal wells is placed
upgradient while a row of injection wells is placed downgradient of a plume. The reverse hydraulic gradient
either stops or reverses the flow of contaminated ground water between the rows of wells while uncontaminated
ground water flows around the affected area.
Special types of well systems have been designed for hydrocarbon recovery because of economic considerations
and the frequency of leaks from underground storage tanks (Canter and Knox, 1985). There are four such systems
and they include systems with 1) one pump and one recovery well, 2) one pump with multiple wells, 3) two
pumps and two recovery wells, and 4) two pumps and one recovery well. Each system takes advantage of the fact
that hydrocarbons are relatively insoluble in water and float on top of the water table. Water-insoluble
compounds other than hydrocarbons may also be recovered using these methods (Nyer, 1985). Systems utilizing
only one pump produce an oil-water mixture which must be separated after recovery; however, the operation and
cost of a single pump system is minimal in comparison to those with multiple pumps (Canter and Knox, 1985). A
single pump system with multiple well points may be the system of choice in soils with low permeability.
Systems utilizing two pumps produce a separate oil and water phase (Canter and Knox, 1985). The system
utilizes a shallow and a deep pump; the deep pump creates a cone of depression while the floating product is
drawn to the center of the cone and recovered by the shallow pump. The shallow and deep pumps are housed in
separate wells in the two pump-two well system whereas both pumps are housed in the same well in the two
pump-one well system. The two pump-one well system requires a large well housing to accomodate two pumps,
produces large quantities of water, and requires continuous and careful monitoring during operation. The draw
down pump of a typical two pump-one well system produces between 100 to more than 300 gallons per minute;
the well casing is 26-inches in diameter and extends 35 ft below the water table (Peterec and Modesitt, 1986).
Better separation of product and water is achieved when there is 30 feet between the depression and recovery
pumps. The disadvantage of the two pump-two well system is the added cost and monitoring of a second well.
The two pump-one well system is the most commonly used. Smith (1986) described a recovery well system for
an aquifer contaminated with hydrocarbons that consisted of five barrier wells located on the downgradient edge
of the dissolved hydrocarbon plume and four product recovery wells (two pump-one well) positioned in the
immediate spill area. The barrier wells removed the dissolved hydrocarbons that were not recovered by the
recovery wells.
In lieu of two pump-one well systems which produce large quantities of water and are expensive to maintain and
operate, Peterec and Modesitt (1986) demonstrated that 60 percent less water is produced and a more effective
cone of depression can be achieved by pumping multiple shallow wells (4-inch diameter) at low flow rates. These
small wells are fitted with pneumatically operated pumps which produce both hydrocarbon and water without
generating an emulsion. In addition, the multiple well system reduces the exposure of uncontaminated areas that
results from dynamic recovery operations using fewer wells. The use of observation wells as recovery wells also
reduces the overall cost of the recovery operation.
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Compounds such as the chlorinated hydrocarbons sink rather than float on top of the water table (Nyer, 1985).
The contaminant must be located by sinking a well directly into the pool of material at the bottom of the aquifer.
Because chlorinated hydrocarbons move in the same direction as the ground water (Keely et al., 1986), wells are
used to force the material in the desired direction (Nyer, 1985). A one-pump system can be used which is
controlled by a conductivity probe. The pump turns on when the pool of chlorinated hydrocarbon reaches the
probe because electricity is not conducted by pure organic material.
The interceptor trenches described in Section B.I. under passive ground water controls may also be active
(pumped). Active interceptor trenches contain vertical removal wells or a perforated, horizontal removal pipe and
are usually backfilled with coarse material for stability (Canter and Knox, 1985). The system must be pumped
continuously to remove the contaminants and prevent discharge downgradient of the trench.
There are many advantages of using well systems for aquifer restoration. Well systems are readily installed and
existing monitoring wells are often incorporated into the design (Canter and Knox, 1985). The design of well
systems is flexible and the cost for installation is often less than that of physical barriers. However, well systems
are expensive and time consuming to operate and maintain. In addition, uncontaminated water is mixed with
contaminated water during withdrawal which results in large volumes of water to be treated at the surface. Years
of treatment may be required to successfully remediate a contaminated aquifer using well systems. Well systems
may be minimally effective or not at all in low permeability formations.
C. Withdrawal and Treatment
Ground water can be treated using a variety of methods after withdrawal from a polluted aquifer. The treatment
of choice depends on the nature of the contamination. Treatment of water contaminated with one compound may
be simple, whereas spills of multiple compounds are complex and may require more than one treatment (Canter
and Knox, 1985). There are numerous methods for physical, chemical and biological treatment of polluted ground
water at the surface. These include biological treatment, carbon adsorption, catalysis, chemical oxidation,
chemical reduction, chemical precipitation, crystallization, density separation, dialysis/electrodialysis, distillation,
evaporation, filtration, flocculation, ion exchange, resin adsorption, reverse osmosis, solvent extraction, stripping,
ultrafiltration, and wet oxidation (Ehrenfeld and Bass, 1984; Nyer, 1985). Processes most applicable to organic
wastes include biological treatment, carbon adsorption, resin adsorption, chemical oxidation, stripping, reverse
osmosis, sedimentation, flotation, and filtration. Many of the processes used have been borrowed from existing
technology concerning water and wastewater treatment.
1. Chemical Treatment
Chemical treatments include such processes as neutralization, precipitation, oxidation-reduction, wet air
oxidation, ozonation, addition of hydrogen peroxide, and coagulation (Ehrenfeld and Bass, 1984). Some general
characteristics of selected chemical processes are given in Table 2-2. Water and wastewater have been treated
extensively using precipitation and coagulation; however, these and other chemical processes are ineffective for
removal of dissolved organic and inorganic materials (Sills et al., 1980). In addition, high concentrations of
chemicals are necessary to treat heavily contaminated material and this produces a large volume of potentially
hazardous sludge. Because of these drawbacks, chemical treatment is often used in conjunction with other
processes.
Oxidation-Reduction-
Oxidation-reduction (redox) reactions may be used to remove some pollutants from ground water (Ehrenfeld and
Bass, 1984). Redox reactions either raise or lower the oxidation state of a substance, thereby decreasing toxicity,
solubility, or rendering the substance more amenable to removal. Currently, there are no demonstrated reduction
reactions for organic compounds; however, many techniques are available for removal of organic compounds by
oxidation.
The reduction of the toxic hexavalent chromium ion to the less harmful trivalent form is an example of a redox
process. The process entails reduction of the hexavalent chromium ion by a reducing agent such as sulfur dioxide,
11
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sodium bisulfite, waste pickling liquor from metal plating, or ferrous sulfate under acidic conditions (pH between
2 and 3). The pH of the solution is usually adjusted with sulfuric acid. The reduction of hexavalent chromium is
rapid (5 to 15 minutes retention time). The resulting trivalent chromic sulfate is treated with alkali and then
precipitated out of solution (Taylor and Qasim, 1983). The precipitate is removed by conventional solids removal
techniques. Before the solution is acceptable for discharge, the pH must be readjusted to near neutrality (pH
between 6 and 9) (Ehrenfeld and Bass, 1984). Before treatment, special precautions must be taken to remove any
cyanide that may be present in the waste; adjustment of a waste containing cyanide to a pH of 2 to 3 may release
hydrogen cyanide. Cyanide may be removed from a waste by chlorination (Ehrenfeld and Bass, 1984).
Chlorination oxidizes the cyanide to nitrogen and bicarbonates and the treatment effectively removes cyanide to
less than 1 ppm. The process involves adjusting the pH between 9 and 11 and adding chlorine gas or sodium
hypochlorite.
Chlorine, hydrogen peroxide, and ozone are chemical oxidants that have been used widely in treatment of
drinking water and industrial wastewater (Nyer, 1985). Ozone and hydrogen peroxide, but not chlorine, are also
used in the treatment of polluted ground water. Chlorine is not used because the oxidation may form chlorinated
compounds, such as chlorinated hydrocarbons, that are extremely undesirable in ground water.
Table 2-2. General Characteristics of Some Chemical Treatment Processes (Law Engineering, 1982)
Process
Removal Efficiency
of Class I and II
Compounds"
Capital and
Operating
Costs
Limitations
Ozone
Hydrogen peroxide
Coagulation
Resin Adsorption13
Ultraviolet
radiation13
Good removal
of I and II
Fair to incomplete
removal of I and
II with extended
contact time
Poor removal of
I and II
Good removal
of some
compounds in I
Good removal of
I and II with
extended time
Moderate to high
Modest capital
and high
operating
Moderate capital
and high
operating
Moderate capital
and high
operating
Moderate to high
Energy
intensive
Incomplete
oxidation of
objectionable
compounds
No information
No information
Expensive
' Class I compounds: xylenes, toluenes, benzene, ethyl benzene, and naphthalene
Class II compounds: ether, alcohols
b Classified as both a chemical and physical treatment
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As an oxidizing agent, hydrogen peroxide is stronger than molecular oxygen but weaker than ozone (Law
Engineering, 1982). However, under alkaline conditions hydrogen peroxide may be a stronger oxidizer than
ozone. Hydrogen peroxide can oxidize aromatic compounds with multiple rings and is readily available (Nyer,
1985). Drawbacks to use of hydrogen peroxide are numerous. Extended contact time for complete oxidation is
required and the final result is often fair to incomplete contaminant removal. The high cost of this treatment
limits its use to small operations (Nyer, 1985). In addition, hydrogen peroxide can impart an after taste in treated
water (Law Engineering, 1982). Therefore, hydrogen peroxide treatment is often used in conjunction with
processes such as ultraviolet radiation.
Ozone reacts with organic compounds to form less harmful, oxidized products (Ehrenfeld and Bass, 1984). The
use of ozone is limited to dilute concentrations (<1%) of oxidizable compounds. Chlorinated hydrocarbons,
alcohols, chlorinated aromatics, pesticides, and cyanides are susceptible to ozonation. Mixing and adequate
contact time of the ozone with the contaminants is crucial because the rate of reaction is limited by mass transfer.
There are numerous drawbacks to ozonation. Ozone is extremely reactive and special equipment that resists
corrosion is required for the operation. In addition, ozone is acutely toxic and appropriate safety measures must
be exercised. Like hydrogen peroxide, the process requires an extended contact time for efficient removal (Nyer,
1985). The capital and operating costs are moderate to high and therefore this technique may be of limited value
in ground water contaminated with organic compounds (Law Engineering, 1982).
The removal efficiency of organic compounds using ozone may be enhanced by combining the treatment with
ultraviolet (UV) radiation (Ehrenfeld and Bass, 1984). The reactor is equipped with UV lamps, positioned in the
water at 3- to 6-inch intervals (Nyer, 1985). The UV light catalyzes the ozonation process. The UV-ozone
process has been demonstrated to rapidly oxidize chlorinated hydrocarbons. Residence times of less than one
minute were required to oxidize 100 ppb trichloroethylene to 0.6 ppb. Ultraviolet radiation may be used singly in
treatment processes. Like hydrogen peroxide and ozone, UV light can be used to destroy organic compounds; in
contrast, water treated by UV light has no odor or aftertaste (Law Engineering, 1982). However, an extended
reaction time is required for good removal and the capital and operating costs are moderate to high. Therefore,
UV light is generally used for small scale operations. Also, UV light cannot penetrate turbid or highly colored
waters.
The Light Activated Reduction of Chemicals (LARC) process uses UV light in a reducing environment to
dehalogenate various organic compounds (Kitchens et al., 1984). The photochemical reaction is conducted in a
solvent, alcohol or water. The medium of choice radically affects the mechanism and efficiency of the process.
Contaminated soil is extracted first and the resulting extract is then photochemically treated. Dehalo- genation
of Arochlor 1260 extracted from soil followed pseudo-first-order kinetics (k = 0.052/min); total dehalogenation
was achieved in 120 minutes. The dehalogenation kinetics of chlordane was pseudo-first-order (k = 0.029/min)
and was complete in 180 minutes. Extraction and treatment of soils contaminated with 1500 mg PCB/kg using the
LARC process is estimated to cost $84.60/ton of soil.
Neutralization--
Neutralization is another type of chemical treatment that is used to adjust the pH of the contaminated ground
water (Ehrenfeld and Bass, 1984). The pH is usually adjusted between 6.0 and 9.0 using acidic or basic materials
such as sulfuric acid, hydrochloric acid, and reagents containing sodium, calcium, or magnesium. Aqueous
samples and a few nonaqueous materials are amenable to neutralization. Safety precautions and appropriate
construction materials must be used because of the corrosive nature of acids and bases and care must be taken to
ensure compatibility of the neutralizing agent and waste. Neutralization, however, does not remove the
contaminant. The process is usually used in conjunction with other processes in a pre- or post-treatment scheme.
Precipitation-
Precipitation is a physical-chemical technique used to decrease the solubility of a particular contaminant by
changing the chemical equilibrium of the waste (Ehrenfeld and Bass, 1984). The technique has been used
routinely to remove inorganic compounds from wastewater (Canter and Knox, 1985). Precipitation may be
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accomplished by adding a reagent that complexes the contaminant, adding a reagent that shifts the chemical
equilibrium of the contaminant, or adjusting the temperature (Ehrenfeld and Bass, 1984). After precipitation, the
insoluble material is removed using conventional methods for solids removal such as sedimentation, filtration, and
centrifugation.
Materials used in precipitation processes include sodium hydroxide, sodium sulfide, calcium hydroxide, inorganic
iron compounds, phosphate salts, and alum. Treatment with sulfide precipitates the most metals, but the resulting
sludge may solubilize when exposed to oxygen (Canter and Knox, 1985). Oxidation of the sulfide complex to
sulfate resolubilizes the metals. The carbonate process is difficult to control. The most widely used precipitation
system employs hydroxides to remove metals; however, the resulting sludge may be viscous and difficult to
dewater. The design for a precipitation system should consider the appropriate precipitating agent, dosage, time
required for flocculation, and method for solids removal (Lee and Ward, 1986). The precipitation process can be
limited by the presence of complexing compounds in the waste such as cyanide or EDTA; these compounds
compete with the precipitant and keep the contaminant in solution (Ehrenfeld and Bass, 1984). Another limitation
is that precipitation produces a sludge that may be hazardous and should be treated accordingly.
Coagulation--
Coagulation is another chemical process with origins in water and wastewater treatment technology. The process
involves the aggregation of molecules of the contaminant to form a floe (Law Engineering, 1982). The floe is
removed by conventional wastewater treatment techniques. The use of coagulation as a remedial treatment is
limited because the process may not effectively remove low molecular weight organic compounds present in trace
amounts. Coagulation is often used together with carbon adsorption for more effective removal of contaminants.
In addition, treatment of high concen-
trations of waste by coagulation is not recommended because of the large volume of potentially hazardous sludge
that is generated (Ghassemi et al., 1980).
Wet Air Oxidation--
Wet air oxidation (WAO) is a form of combustion that involves the addition of pressurized air to liquid wastes at
high temperatures (Ehrenfeld and Bass, 1984). The process is suitable for the treatment of high concen-
trations of wastes with high removal efficiencies for low molecular weight organics and oily substances
(Ghassemi et al., 1980). WAO requires highly skilled personnel, specialized construction materials, and cooling
water.
2. Physical Treatment
Physical treatment methodologies include activated carbon adsorption, density separation, filtration, incineration,
reverse osmosis, ion exchange, and air and steam stripping (Ehrenfeld and Bass, 1984).
Carbon Adsorption-
The most widely used treatment for liquid contaminants is activated carbon (Sills et al., 1980). Activated carbon
is generally used to remove dissolved organic contaminants that are not removed by biological treatment. Some
inorganic compounds such as antimony, arsenic, bismuth, chromium, tin, silver, mercury, and cobalt are partially
removed (Ehrenfeld and Bass, 1984). The method is versatile because complex mixtures of organic compounds
and volatile compounds can be treated (Hall and Mumford, 1985). The process involves passing an aqueous
stream of waste through particles of granular activated carbon (GAC) or powdered activated carbon (PAC).
The contaminants are trapped and held on the carbon particles by physical processes (Nyer, 1985). The adsorbent
nature of the carbon is a result of a high internal surface area which is independent of particle size. Granular
carbon is widely used because it can be thermally regenerated, a process which also destroys the sorbed pollutants
(Ehrenfeld and Bass, 1985). The carbon is usually thermally regenerated in a multiple hearth furnace but steam
treatment, solvent extraction, and biological treatment are also used. Adsorption of organic material using
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granular carbon employs moving or fixed bed reactors. Saturated adsorbent must be replaced or regenerated
periodically. In contrast, powdered activated carbon is added to the aqueous waste and mixed for desired periods
of time. After sufficient contact time, the powdered carbon is removed. The spent carbon is usually not
regenerated. Powdered carbon may be used in the clarification step of biological treatment or added to aerated
sludge. The spent particles settle out with the sludge in the latter application and may improve the settling
properties of the sludge.
In conventional wastewater systems, carbon adsorption is generally used for wastes containing 1 percent or less
adsorbable contaminants; however, the technique is often used for more concentrated wastes in remedial action
(Ehrenfeld and Bass, 1984). The amount sorbed depends on the equilibrium of the solute between the carbon
particles and solution (Canter and Knox, 1985). The affinity of the carbon for the organic substance depends on
the 1) solubility of the organic compound, 2) solution pH, 3) molecular structure of the organic compound, 4)
temperature, and 5) surface characteristics of the adsorbent.
Proper design of an activated carbon unit for treatment of contaminated ground water involves selection of carbon
with appropriate physical properties and determining the carbon dosage and contact time required for efficient
removal. Equilibrium adsorption isotherms are used to determine whether the contaminant can be effectively
removed, the adsorption capacity of the carbon for the contaminant, and the amount of carbon required (Canter
and Knox, 1985). Estimates of treatability can be made from the adsorption isotherms; however, column studies
in the laboratory should be conducted to determine contact time, bed depth, pretreatment requirements, head loss
characteristics, and carbon dosage (Canter and Knox, 1985). The breakthrough curve for the carbon column
should also be determined (Ehrenfeld and Bass, 1984). An example of such a study was reported by Chrobak et
al. (1985) in which pilot plant experiments were used to predict the performance of GAC adsorption in removal of
chlorinated solvents.
Even with careful design based on good data, carbon adsorption is affected by fluctuation in flow rates and the
composition of the waste (Ehrenfeld and Bass, 1984). Therefore, systems are generally designed to compensate
for these variations. In addition, the carbon must be replaced or regenerated before binding sites are saturated and
adsorption efficiency declines. Hall and Mumford (1985) reported that the life expectancy of the carbon bed is
affected by the type and concentration of the pollutants, column size, and volume of water treated. Microbial
growth may occur in the carbon system which may improve adsorption by removing contaminants or hinder the
process with biofouling and/or production of objectional odors.
Granulated activated carbon (GAC) is well suited for treating ground water polluted with contaminants in the 500
ppb or less range (Stenzel and Gupta, 1983). Such conditions are often found in contaminated aquifers tapped for
drinking water. GAC has been reported to remove contaminants in this concentration range to nondetectable
levels. In addition, GAC is an excellent choice for remedial action of ground water contaminated with higher
concentrations (ppm) of pollutants. Such contamination regimes may result from percolation of leachate from
lagoons or waste sites into shallow aquifers. GAC may be the preferred treatment over air stripping in this case
because stripping releases hazardous emissions and/or may not remove contaminants to acceptable levels.
Use of GAC in remedial action is cost-effective because of the regenerative capacity of the carbon (Stenzel and
Gupta, 1983). The overall cost of treatment depends on flow rate, type and concentration of contaminants, and
site requirements (Canter and Knox, 1985). Treatment of wastes in the ppm range may cost between $0.48 to
$2.52/1000 gallons whereas $0.22 to $0.55/1000 gallons was reported for concentrations in the ppb range.
Resin Adsorption--
Resin adsorption is a physical-chemical process in which dissolved substances are removed from solution by
synthetic or natural materials (Law Engineering, 1982). The resins may be ionic or nonionic; the nonionic
materials form stronger bonds which are harder to reverse. Therefore, weak ionic or nonionic resins are more
widely used. Synthetic resin adsorbs dissolved substances by trapping and holding the solutes in its molecular
structure (Nielsen, 1983). In comparison to activated carbon, resin adsorption is more selective for some
compounds; non-ionic resin polymers readily sorb low molecular weight lipophilic substances such as the alkyl
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benzenes (Law Engineering, 1982). In addition, the rate of resin fouling is slower than that of carbon adsorbants.
After saturating the binding sites, the resin can be regenerated chemically at relatively low cost (Nielsen, 1983).
However, use of resin adsorption may be limited to small operations because of the moderate capital and high
operating costs (Law Engineering, 1982).
Ion exchange--
Ion exchange is another type of adsorption process which has been demonstrated for removal of most inorganic
dissolved salts, some organic dissolved salts, and some low molecular weight lipophilic compounds (Ehrenfeld
and Bass, 1984; Law Engineering, 1982). In the past, this process has been used to remove unwanted ionic
species from waste streams (Ehrenfeld and Bass, 1984); however, the use of weak ionic resins has generated
interest in using the process to remove dissolved organic contaminants (Law Engineering, 1982). Ion exchange
can be cost-effective because the treatment bed can be regenerated (Ehrenfeld and Bass, 1984). For regeneration
to be practical, waste streams should contain 2500 mg of total dissolved solids/liter or less. Cost effectiveness is
reduced and the application of ion exchange is limited for more concentrated wastes.
Stripping-
S tripping with air or steam is used to remove dissolved volatile organic compounds from wastewaters (Ehrenfeld
and Bass, 1984). Air stripping involves transferring a dissolved substance from the liquid to the gas phase
whereas steam stripping is essentially a distillation process in which the volatile pollutants are removed from the
waste as the distillate. Removal of trihalomethane and trichlorethylene has been demonstrated for air stripping.
Volatile organic compounds, ammonia, hydrogen sulfide, and water-insoluble compounds such as the chlorinated
hydrocarbons, are amenable to steam stripping. High removal efficiencies using either air or steam stripping can
be achieved. Removal of trichloroethylene from ground water by air stripping exceeded 99 percent whereas 10 to
99 percent of some volatile organic compounds was removed by steam stripping. The stripping potential of a
particular compound can be predicted using Henry's Law constant (Canter and Knox, 1985). Compounds with
high Henry's Law constants are more easily removed. The driving force of the removal is the concentration
differential between the liquid and air phases; the rate of mass transfer is also important.
Air stripping can be accomplished using aeration tanks, cascade aerators, spray basins, or packed towers; the latter
two methods are more economically and technically feasible (Nyer, 1985). The packed tower is the more
commonly used method for remedial action on ground water because of the high interfacial area and high air-to-
water volume ratios (Canter and Knox, 1985; Nyer, 1985). The system operates in a countercurrent fashion by
which the aqueous waste flows downward through a tower of packing material while the air current is forced
upward. The stripped water is collected at the bottom of the tower for proper disposal or treatment and the
volatiles are emitted from the top with the air.
Factors that control stripping efficiency include tower height and diameter, the air-to-water ratio, and the
temperature (Canter and Knox, 1985; Nyer, 1985). Elevated temperatures will increase the rate of stripping in
addition to removing compounds that are not stripped at lower temperatures. Packed towers include an aeration
system, devices to insure that mass transfer occurs optimally, and packing material in the tower shell. The
packing material should create turbulence and provide a large surface area for air-water contact; corrosion
resistance, weight, and cost of the packing material should also be considered. Polypropylene is often used as the
packing material because the plastic is chemically inert, inexpensive, lightweight, and strong. Design of an air
stripping tower should be preceded by pilot studies that incorporate site specific characteristics (Stover and
Kincannon, 1983).
The obvious limitation in the use of air stripping is the release of hazardous pollutants into the atmosphere (Nyer,
1985). However, the pollutants that are released into the atmosphere are less concentrated than those in the
wastewater because of dilution. In addition, some compounds such as tri- and tetrachloroethylene that are
released in the atmosphere are subject to photolysis by the sun. Finally, hazardous emissions from stripping
towers may be treated by activated carbon or incineration. A combination of air stripping with activated carbon
may be less expensive than carbon treatment alone. Stripping towers are also subject to freezing, biofouling, and
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clogging. Estimates for stripping trichloroethylene with a 90 percent efficiency ranged from $0.09 to $0.90/1000
gallons (Canter and Knox, 1985).
Incineration--
Incineration is a conventional and demonstrated technique used to destroy organic wastes (Ehrenfeld and Bass,
1984). The process theoretically involves the complete destruction of organic materials to inorganic species such
as carbon dioxide, water, sulfur dioxide, oxides of nitrogen, and hydrochloric acid by raising the temperature of
the waste in the presence of oxygen; temperatures ranging from 815 to 1094°C are employed for the oxidation
(Nyer, 1985). Rotary kilns are most commonly used to incinerate mixtures of wastes with solid residues but
multiple hearth, fluidized bed, and liquid injection incinerators are also available (Ehrenfeld and Bass, 1984).
Multiple hearth incinerators accomodate liquids and gases in addition to solids and sludges, whereas liquid
injection incineration is limited to liquids and slurries that can be pumped. The fluidized bed incinerator is suited
for solids, liquids, and gases.
The suitability of a waste for incineration depends on multiple characteristics such as moisture content, volatile
substances content, specific gravity, metal content, flammability, reactivity, and toxicity (Ehrenfeld and Bass,
1984). A self-sustained burn can be achieved with aqueous wastes containing 20 percent or more organic
material; however less concentrated wastes will require auxiliary fuel (Nyer, 1985). Ground water is rarely
polluted with 20 percent or more organic wastes.
A federally regulated incinerator must achieve an afterburner temperature of 1200°C and a two second dwell time
and requires three percent excess oxygen (Ehrenfeld and Bass, 1984). Emission control equipment to remove
sulfur dioxide, hydrochloric acid, and products of incomplete incineration should also be included when
appropriate. Before a full scale incinerator is designed, pilot tests are conducted to determine the residence time,
temperature, destruction efficiencies, ash residue, and gaseous effluent for a particular waste.
Incineration may be economically feasible for concentrated waste that can sustain a burn (Nyer, 1985).
Incineration of small quantities of wastes may not be economically feasible unless the treatment facilities are
nearby or mobile incineration units are available. Limitations other than cost include potential air pollution and
the substantial upkeep that is required.
Design of a mobile incineration unit for the EPA was started in 1976 to provide a more cost-effective process
(Freestone and Brugger, 1980). A rotary kiln design was chosen to accommodate a variety of organic wastes. A
trial burn using the mobile unit was conducted and destruction efficiencies of 199.99 percent were reported for
carbon tetrachloride, chlorinated benzenes, and Aroclor 1260 (PCB) at feed rates of 70,95, and 50 Ib/hr,
respectively (Yezzi et al., 1984). The EPA has also been involved in the development and feasibility studies of
another mobile incineration system, a High-Temperature Fluid-Wall (HTFW) reactor (Hornig, 1984). The
removal efficiency for a 1 percent spike of Aroclor 1242 in soil was 99.9997 percent.
Reverse Osmosis-
Reverse osmosis (RO) is a filtration process by which inorganic salts and some organic substances (with
molecular weights >300 g/mol) are removed from solution by passing aqueous wastes through a semi-permeable
membrane under pressure (Ehrenfeld and Bass, 1984). The inorganic and some organic substances are
concentrated on one side of the membrane while water passes through (Nyer, 1985). The osmotic pressure of the
waste solution is counteracted by the applied pressure so that the solute can be concentrated (Ehrenfeld and Bass,
1984). Pressures between 200 to 400 psi are required.
The most commonly used membranes in RO include cellulose acetate, cellulose triacetate, polyamides, and
polysulfones; however, many other membranes are in the developmental stages (Ehrenfeld and Bass, 1984).
Cellulose acetate is most commonly used; however, the other membranes may be better suited for a particular
waste depending on the pH, temperature, and nature of the contaminants. A polyether polysulfone membrane was
reported to remove at least 90 percent of the benzene, hexane, and many chlorinated solvents in leachate at
concentrations in the low to middle ppb range (Whittaker, 1984).
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In general, removal efficiencies of total dissolved solids (up to 50,000 mg/L) from waste streams is greater than
90 percent (Ehrenfeld and Bass, 1984). Organic compounds of low molecular weight are not removed by RO
(Nyer, 1985); the removal efficiency of organic compounds usually decreases with an increase in polarity and
hydrogen bonding with the filter (Ghassemi et al., 1980). Removal efficiencies for oily substances and most
heavy metals by RO are greater than 90 percent.
The limitations of wastewater treatment using RO are related to membrane fouling. The removal efficiency of the
membrane can be compromised by suspended solids, biological growth, strong oxidizers, extremes in pH, and
high concentrations of substances such as phenols, calcium, silica, sulfate, and aluminum (Ghassemi, et al., 1980).
Pretreatment may be necessary when the waste stream characteristics are variable (Ehrenfeld and Bass, 1984). In
addition, RO is expensive because of the high pressure required to drive the process (Nyer, 1985).
Density Separation--
Density separation encompasses the processes of sedimentation and dissolved air flotation (Ehrenfeld and Bass,
1984). Sedimentation involves the removal of suspended solids that settle out by gravity and is often used after
chemical precipitation and biological treatment using activated sludge. The process can be enhanced by the
addition of flocculants. Flocculation is a two-step process (Nyer, 1985). The first step is to neutralize the charge
(usually negative) on the suspended particles so that flocculation can occur. Neutralization may be achieved using
inorganic coagulants such as lime, alum, and ferric chloride. Other agents may be used when the treated water
will be used for drinking water supplies. The second step in flocculation is to coalesce the smaller particles to
produce larger particles, or floes. High concentrations of small particles usually coalesce with gentle mixing.
However, gentle mixing may not be sufficient to flocculate low concentrations of suspended particles. In this
instance, special equipment and the addition of organic flocculating agents may be required. Sedimentation is
attractive because of the low cost (Ehrenfeld and Bass, 1984).
Dissolved air flotation is used to remove insoluble particles of pollutants or nonaqueous phases of compounds
such as hydrocarbons from waste mixtures (Ehrenfeld and Bass, 1984). The waste mixture is first aerated at high
pressures and then transferred into a tank at atmospheric pressure. The change in pressure creates fine air bubbles
that rise and carry the insoluble particles and nonaqueous phases of compounds to the surface. The floating
material is then skimmed off the surface. The flotation process can be enhanced by the addition of surfactants.
Density separation is of questionable use in removal of hazardous wastes because of the sludge that results from
both sedimentation and flotation processes (Ehrenfeld and Bass, 1984). In addition, complete removal of the
hazardous materials may not be achieved. The dissolved air flotation process may also produce hazardous
emissions which should be treated appropriately.
Filtration--
Filtration is an effective method that separates low levels of solids from an aqueous phase by passing the waste
mixture through a semi-permeable medium (Ehrenfeld and Bass, 1984). In addition to producing a purified
filtrate, filtration also decreases the volume or dewaters the waste mixture to produce a sludge concentrate. There
are three types of filtration units: granular media filters, rotary drum vacuum filters, and filter presses. Granular
media, such as sand, is used to filter suspended solids by gravity or auxiliary pressure. The system is most
efficient at filtering aqueous wastes, whereas filtration efficiencies for nonaqueous liquid wastes, slurries, and
sludges are moderate to low. The filter bed is regenerated by backwashing and removing solids by flocculation or
sedimentation. This type of filtration removes suspended solids down to the 1 to 10 mg/L range.
The rotary drum vacuum filters are constructed by stretching fabric or wire mesh over a drum and small roller
(Ehrenfeld and Bass, 1984). The drum is partially submerged in the aqueous mixture which is drawn into the
drum by a vacuum. The filtrate is collected and the trapped solids are removed from the filter. The rotary drum
vacuum filter can be used for slurries and sludges in addition to aqueous and nonaqueous liquid wastes; sludges
can be dewatered by 60 to 90 percent, but high concentrations of suspended solids still remain in the filtrate.
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The filter press is constructed with a series of plates and screens which trap solids by squeezing the water out of
the waste mixture (Ehrenfeld and Bass, 1984). This type of filtration system is often used to dewater viscous
sludges in addition to slurries, and aqueous and nonaqueous liquid wastes. Sludges can be dewatered by 50 to 85
percent, but high concentrations of suspended solids still remain in the filtrate.
In general, the efficiency of filtration can be reduced by clogging the filter bed or fabric medium. In addition,
filtration of hazardous wastes results in trapped suspended solids or dewatered sludges that are also hazardous.
Dissolved hazardous compounds may not be removed by the filtration process and thus require further treatment.
3. Biological Treatment
Ground water can be withdrawn and treated by conventional biological wastewater treatment processes.
Treatment processes used to treat contaminated ground water and leachate from hazardous wastes include: 1)
suspended growth processes such as activated sludge, lagoons, waste stabilization ponds and fluidized bed
reactors, and 2) fixed film processes such as trickling filters, rotating biological discs, sequencing batch reactors
and others. The wastewater can be treated on site by one of these processes or off site at a municipal or
commercial treatment plant. A short discussion of each process follows with regards to hazardous waste
management. More detailed discussions of these processes are presented in Johnson, 1978; J. R. B. Associates,
1982; Ehrenfeld and Bass, 1984; Shuckrow et al., 1980; Nyer, 1985; and Canter and Knox, 1985.
Biological Wastewater Treatment Processes--
The most commonly used municipal wastewater treatment process, activated sludge, has a number of advantages
for treatment of contaminated ground waters. The units include an aeration basin, a clarifier, and sludge recycle.
Following treatment in the aeration basin, a portion of the sludge collected in the clarifier is recycled. The
recycling process allows an acclimated microbial population to build up in the system, hence the name activated
sludge. The settled sludge can adsorb heavy metals and some organics which may cause the sludge to be
considered a hazardous waste (Shuckrow et al., 1980). However, some compounds are removed by volatilization
during the aeration step (Eckenfelder et al., 1985). Activated sludge treatment can reduce the soluble biochemical
oxygen demand (BOD) to less than 10 mg/L and the total BOD, including suspended solids, to less than 30 mg/L.
The retention time is short and consequently the process is sensitive to toxic and hydraulic shocks. A survey of 92
industrial wastewater streams conducted by the EPA reported mean BOD removal efficiencies of 86 percent
(Ehrenfeld and Bass, 1984). Specific organic compounds can be degraded to low levels; effluent levels of phenol
as low as 0.02 mg/L have been reported (Nyer, 1985).
The sequencing batch reactor is an application of the activated sludge process which may be used to treat
hazardous waste leachates (Ying et al., 1986). The process involves five steps per cycle: 1) fill-the wastewater is
drawn into the vessel where some of the activated sludge from the previous cycle remains, 2) react-aeration and
mixing occur, 3) settle-clarification occurs in this step, 4) draw-the supernatant is withdrawn, and 5) idle-the
system remains idle until the next cycle is initiated. Treatment of hazardous waste with the sequencing batch
reactor may be advantageous because the process is more complete and flexible than other treatment technologies
and can provide intermittent treatment; in addition, the same tank can be used for both treatment and clarification.
Up to 90 percent of the TOC from a hazardous waste leachate was removed under a 24-hour cycle with a 10-day
retention time (Ying et al., 1986).
Like activated sludge, surface impoundments such as aerobic lagoons, facultative lagoons, anaerobic lagoons, and
waste stabilization ponds rely on suspended microbial populations to degrade organic material; unlike activated
sludge, the biomass is not recycled (Wilkinson et al., 1978; Ehrenfeld and Bass, 1984). Even though the processes
typically require less energy and supervision than activated sludge, the operational controls are not as flexible
(Shuckrow et al., 1980). The retention time of waste in a surface impoundment is often on the order of weeks
whereas that of activated sludge may take a few hours. In general, surface impoundments are quite large, and
their size allows for dilution and buffers fluctuations in organic load. Aerobic lagoons are aerated mechanically or
by diffusion to increase the degradation rate of organic material and mix the system (Johnson, 1978; Eckenfelder
et al., 1985; Ehrenfeld and Bass, 1984). Organic material is degraded aerobically at the surface and anaerobically
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near the bottom of facultative lagoons (Ehrenfeld and Bass, 1984). Because aeration is not forced or used,
facultative and anaerobic lagoons offer the advantage of easy operation and low cost (Shuckrow et al., 1980);
however, anaerobic processes result in incomplete degradation of organic compounds and hence low quality
effluent. Facultative lagoons can also tolerate higher organic loading than aerobic lagoons and both facultative
and anaerobic lagoons may generate noxious odors (Eckenfelder et al., 1985). Anaerobic processes are enhanced
in anaerobic lagoons by low surface to volume ratios (Ehrenfeld and Bass, 1984). Waste stabilization ponds are
lagoons that are aerated by natural processes such as wind and photosynthesis (Ehrenfeld and Bass, 1984). The
ponds are principally a polishing technique for low organic waste waters (Johnson, 1978). The ponds are usually
0.3 to 0.6 m in depth (Ehrenfeld and Bass, 1984). Removal efficiencies for surface impoundments are in the
range of 60 to 90 percent. They are sensitive to shock loadings of toxic chemicals and fluctuations in temperature.
Fluidized bed reactors are filled with materials such as sand or coal that are suspended by wastewater which flows
upward through the material (Wilkinson et al., 1978). The particles are colonized by a dense growth of
microorganisms which rapidly degrade organic material present in the waste stream.
In the fixed film process, wastewater is passed over a surface colonized with microorganisms; the attached biofilm
degrades the organic material. The original fixed film process, the trickling filter, uses a solid medium such as
rock or plastic, as the surface for microbial attachment (Canter and Knox, 1985). Trickling filters can remove
from 60 to 85 percent of the BOD (Eckenfelder et al., 1985). Suspended or colloidal organics can be treated, and
the process is usually limited to low organic loadings (J. R. B. Associates, 1982).
Trickling filters that are operated anaerobically are known as anaerobic filters (Eckenfelder et al., 1985). The
anaerobic filter process can tolerate high loading rates. Low pH and inorganics such as sodium, sulfate, and
heavy metals may inhibit methanogenesis and toxic organics may also be a problem (Rittman and Kobayshi,
1982). The biological tower is another variation of the trickling filter. The tower is packed with a colonizable
surface which may reach a height of 16 to 20 feet (Canter and Knox, 1985). The process operates in a
countercurrent mode; contaminated water is sprayed on the top of the tower as air is pulled from the bottom.
A rotating biological disc is similar in concept to trickling filters. Discs or drums which are coated with a biofilm
are partially submerged and rotated through the wastewater (Wilkinson et al., 1978; Ehrenfeld and Bass, 1984).
Rotation of the discs aerates the attached biofilm. The process is sensitive to shock loading and temperature
fluctuations, but is otherwise moderately reliable (J. R. B. Associates, 1982). In comparison to activated sludge,
rotating biological discs require less energy and are easier to operate, but are similar in effectiveness (Eckenfelder
et al., 1985). Clarification may be required before and after treatment of wastewater with rotating biological discs
(Ehrenfeld and Bass, 1984).
Biological wastewater treatment processes may be used to treat the following classes of organic compounds
typically found in ground water: alcohols, organic acids, aldehydes, ketones, quinones, amines, amides,
carbohydrates, esters, some ethers, phenolics, and some aromatics (Absalon and Hockenbury, 1983; Eckenfelder
et al., 1985). Compounds that may be difficult to treat with biological wastewater processes include halocarbons,
high molecular weight polynuclear aromatics, pesticides, and organometals (Ehrenfeld and Bass, 1984).
Examples of Withdrawal and Biological Treatment--
Biological wastewater treatment processes have been used in remedial action at several hazardous and
nonhazardous waste sites. Adequate treatment of leachates from recent municipal refuse landfills that contained
high levels of free fatty acids was achieved by biological treatment (Chian, 1977). Leachates from older landfills
may be more amenable to physical-chemical treatment processes. In a pilot study, Stover and Kincannon (1982)
were able to decontaminate ground water from a hazardous waste site using activated sludge. The batch activated
sludge pilot system was seeded with organisms that were acclimated to the same compounds found in the
contaminated ground water - phenols, cresols, dichlorobenzenes, and others. Following acclimation and
stabilization of the batch activated sludge for three weeks, the organisms were able to reduce the total phenols,
TOC, BOD, and chemical oxygen demand by 80 percent or more within 24 hours.
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Combinations of Biological Treatment With Other Processes--
Combinations of conventional biological wastewater treatment with other water treatment processess such as
granular activated carbon (GAC), air stripping, and addition of acclimated bacteria have also been successful.
Feasibility studies on decontamination of leachate from the Ott/Story hazardous waste site in Muskegon, Michigan
have combined activated sludge and GAC adsorption to remove various halogenated aliphatics, benzene, and
toluene (Shuckrow and Pajak, 1981). Initial attempts to acclimate an activated sludge culture to the organic
contaminants were minimally successful and the addition of a commercial microbial culture was not effective;
however, a combination of GAC adsorption followed by activated sludge removed greater than 95 percent of the
TOC. The activated sludge organisms removed the organics that were not sorbed in the GAC treatment.
Treatment efficiencies were greater than 75 percent as long as the removal efficiency of the GAC was high.
Anaerobic treatment combined with GAC was less effective than aerobic treatment (activated sludge) and suffered
the same decline as the aerobic process when the GAC was saturated (Shuckrow et al., 1980). Removal of TOC
by a treatment train using GAC followed by anaerobic and then aerobic biological treatments was also tested
(James et al., 1981). The biological treatment steps were not necessary when the GAC sorption sites were not
significantly saturated; however, the biological treatment increased the removal efficiency of the treatment train
as more GAC sorption sites became saturated. The TOC removal efficiency of GAC combined with anaerobic
and aerobic treatment was less than that of GAC and aerobic treatment only.
Josephson (1983) reported that a combination of powdered activated carbon and activated sludge was used to treat
ground water contaminated with hydrocarbons, pesticides, and other organics. A removal efficiency of 95-99
percent was achieved for the COD, total nitrogen, and various organics.
Air stripping followed by biological treatment was used in a pilot study to treat ground water contaminated with
trichloroethylene, freon, 1,2-dichloroethene, toluene, ethyl benzene, xylenes, vinyl chloride, acetone, isopropanol,
o-dichlorobenzene, 1,1-dichloroethylene, and 1,2-dichlorophenol (Schaezler and St. Clair, 1985). Air stripping
could remove all of the organics except for the nonvolatile compounds isopropanol and acetone; however, a
treatability study indicated that the nonvolatile compounds were biodegradable. Hence, a treatment system
utilizing both air stripping and biological treatment was recommended for this site.
Nyer and Sauer (1984) described the cleanup of shallow ground water from a Gulf Coast hazardous waste site.
The saline aquifer was contaminated with 400 mg/L phenol and other organics which resulted in a TOC
concentration of 1,300 mg/L. Several options that were considered for treatment are reported with estimated costs
per gallon in parenthesis: pond evaporation with oxygen and nutrient addition ($0.028); deep well injection
($0.183); solidification/adsorption to concentrate the liquid and then adsorb the material to trench backfill
($0.085); granular activated carbon adsorption ($0.058); biological treatment ($0.005).
A feasibility study indicated that the organics were biodegradable, but carbon adsorption would be required as a
polishing technique (Nyer and Sauer, 1984). The overall treatment system would consist of: 1) pH adjustment, 2)
chemical addition, 3) biological treatment with two aeration basins, a clarifier, and a fixed activated sludge
treatment system (FAST), 4) filtration through a dual media filter, and 5) carbon adsorption. The FAST system
uses particles of plastic colonized by bacteria in a well mixed tank. The system is essentially a hybrid of activated
sludge and fixed film processes. The biological system was seeded with activated sludge from a refinery that
treated ballast from oil tankers. Results from the pilot plant, which included the aeration plant and clarifier only,
indicated that the TOC was reduced by 70 percent. Addition of the FAST system, the dual media filter, and the
carbon adsorption unit to the treatment train reduced the concentration of TOC from 1,300 to 18 mg/L (98
percent) (Nyer, 1985).
Addition of mutant bacteria to a sequencing batch reactor, a process patented in 1985 by Colaruotolo et al.
(1985b), was used to treat leachate collected from the Hyde Park Landfill in Niagara Falls, New York (Ying et al.,
1986). The leachate contained chlorinated organics, phenol, and benzoic acid. A consortium of microorganisms
that could degrade most of the contaminants was isolated from the leachate; however, degradation of the
pollutants by bacterial strains in the consortium was variable (Sojka et al., 1986). By genetic manipulation,
organisms in the consortium that could degrade the remaining compounds were found. Tests with bench scale
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sequencing batch reactors and also pilot-plant scale units indicated that the TOC was reduced by 85 percent or
greater and that individual contaminants were generally reduced by 95 percent or greater. The biomass yield was
0.64 mg/mg feed TOC. Amendments of nitrogen and phosphorous did not improve treatment over the addition of
only nitrogen. Cost savings for the biological treatment over the existing carbon adsorption system were
estimated to range from $538,000 up to $783,000.
Bartha (1986) suggested that inoculation of microorganisms into a wastewater treatment process be judged with
caution. Inoculation may be useful for startup, disruptions, for certain xenobiotics that cannot be degraded by the
natural flora, or when the added organisms cannot sustain themselves.
D. In Situ Physical and Chemical Treatment
In situ treatment techniques are similar to methods used to treat waste after withdrawal or excavation and methods
used to solidify or stabilize a waste for safer transportation and disposal (Ehrenfeld and Bass, 1984). In situ
remedial action is a direct application of these techniques to the waste in place. Treatment in place avoids the
cost of withdrawal or excavation and disposal after the treatment; the concept is also desirable because there is
little or no visible disturbance of the environment. Also, large quantities of soil can be treated at once (Truett et
al., 1982). In addition, in situ treatment may be more effective than withdrawal and treatment; under certain
hydrogeological conditions, incomplete recovery of the pollutants occurs using conventional pumping techniques
(Ehrenfeld and Bass, 1984). However, in situ techniques may have to be used with physical containment
measures to prevent further migration of the pollutants (Truett et al., 1982).
In situ treatment must be preceded by a thorough hydrogeological investigation of the site and characterization of
the waste (Ehrenfeld and Bass, 1984). Physical and chemical treatments are usually applied to homogeneous
wastes for optimum results and the treatment is waste-specific. The major limitation of in situ physical/chemical
remediation is the potential to create additional pollution or volatilize toxic chemicals during the process
(Ehrenfeld and Bass, 1984).
Chemical treatment involves neutralizing, precipitating, oxidizing or reducing, or destroying the contaminants by
injecting the reactive material into the waste (Ehrenfeld and Bass, 1984). The chemical is added into injection
wells to treat plumes of leachate and ground water, therefore, the extent of the plume must be well defined for
thorough treatment. The reactive agent is added directly to the waste contained in surface impoundments or
landfills.
In situ chemical treatment requires that the waste be homogeneous and that the characteristics and concentrations
of the contaminants are known. Examples of chemical treatment are the oxidation of cyanide with strong
oxidizing agents such as sodium hypochlorite, precipitation of metals with alkali agents or sulfides, and the
precipitation of hexavalent chromium using reducing agents (Ehrenfeld and Bass, 1984). The addition of
activated carbon directly to a surface impoundment or landfill is an example of in situ physical treatment.
A plume of leachate or contaminated ground water can be directly treated using permeable treatment beds or by
chemical injection (Ehrenfeld and Bass, 1984). Permeable treatment beds are constructed by excavating trenches
downgradient of the plume. The trenches are excavated through the aquifer and into the consolidated zone. The
trench is then capped to prevent percolation of rain or runoff into the treatment area. The width of the trench
depends on the permeability of the chosen fill material, the velocity of ground water flow, and the time required to
react the contaminants with the fill material. The technique is amenable to shallow aquifers only because of the
construction design. The trenches are then filled with permeable materials which neutralize or precipitate the
contaminants. Materials such as limestone or crushed shell, activated carbon, glauconitic green sand (actually a
clay with high metal adsorbing capacity), zeolites, and synthetic ion exchange resins are used to fill the trench.
Most of these materials are somewhat effective in removing heavy metals. Limestone can be used to neutralize
acidic plumes while lipophilic organic compounds can be removed by activated carbon. The effectiveness of
permeable treatment beds is short-lived because the fill material becomes plugged. The plume may escape
treatment because the ground water is diverted around the trench or is channeled through the fill material. In
addition, changes in the composition of the plume may affect the contact time and reduce removal efficiencies.
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In situ chemical treatment of a formaldehyde spill was conducted using an alkaline solution of hydrogen peroxide
(Sikes et al., 1984). About 15,000 square ft of formaldehyde-contaminated soil was first treated with a buffered
solution of sodium hydroxide and soda ash (pH 9.5) and then amended in two separate treatments with a 5 percent
solution of hydrogen peroxide. The oxidation rate of formaldehyde by hydrogen peroxide had been observed to
proceed faster at alkaline pH in laboratory tests. The formaldehyde concentration was reduced from 30,000 ppm
to a range of 500 to 1000 ppm after both treatments. The in situ chemical treatment with hydrogen peroxide was
then followed by biological treatment.
Removal of iron and manganese from ground water by the Vyredox method is an example of an in situ chemical
treatment (Hallberg and Martinell, 1976; Zienkiewicz, 1985). The process is used to remove these metals before
the ground water is used for industrial purposes or drinking water. The iron and manganese are precipitated out of
solution in an oxidized state by perfusing the aquifer with oxygen-rich water. Iron and manganese oxidizing
bacteria may be involved in the oxidation reactions; however, microbial participation is not well defined. The
oxidizing conditions maintain high Eh and pH values and the metals precipitate in the strata before reaching the
production well.
Contaminants may be stabilized in the subsurface by polymerization reactions. Williams (1982) reported that an
underground spill of an acrylate monomer was polymerized in place by injecting a catalyst and activator into the
area of contamination. The acrylate monomer is a moderately water soluble liquid that is less dense than water.
Two treatments with a catalyst and an activator four days apart polymerized 85-90 percent of the liquid monomer
into a solid.
Solution mining, or extraction, is another type of in situ physical treatment process (Ehrenfeld and Bass, 1984).
Chemical processing and mining industries have been using chemical extraction techniques for years. In situ
treatment of hazardous waste involves application of a solvent to the waste in place and then withdrawal of the
elutriate using a well point system. The type of compounds removed by this process depends on the
characteristics of solvent used and the contaminants present. Solvents used in the process include water, acids,
ammonia, and/or chelating agents. Laboratory tests should be conducted to determine the extraction efficiency of
the solvent for the target contaminants. Problems that may be encountered include the incompatibility of the
solvent with some compounds in the waste. Also, injection of the solvent into the subsurface may create
additional contamination. The elutriate is also considered hazardous and will require treatment or proper disposal.
The process is largely undemonstrated for in situ remediation: however, metal extraction from some hazardous
materials has been successful in laboratory studies. Ellis et al. (1984) conducted both column and batch
experiments and found 90 percent or greater extraction efficiencies for aromatic and intermediate molecular
weight aliphatic hydrocarbons, polychlorinated biphenyl mixtures, and chlorinated phenol mixtures from soil by
water washing and surfactant solutions.
In situ vitrification is a physical treatment process in which waste and soil are heated to melting and then cooled
to form a glassy, solid material that resembles obsidian (Ehrenfeld and Bass, 1984; Battelle Labs, 1985). Battelle-
Northwest has developed a vitrification process which uses electric-joule heating. An electric current is passed
through the waste from electrodes that are placed around the site; the waste is melted from the top down. The
process is amenable to inorganic and nonvolatile organic compounds/soil mixtures. Treatment of some volatile
organics may release toxic fumes; however, Battelle-Northwest has devised a covering hood to trap any escaping
gases from the treatment area. The trapped gases are then treated in an off-gas processing system.
The resulting vitrified material is about twice as strong as unreinforced concrete with permeabilities that
approximate those of Pyrex glass (Battelle Labs, 1985). In addition, the glass-like material should not pose any
environmental risks for thousands of years. The disadvantages of the process are related to energy consumption.
About 2000 kw/m3 and a melting temperature of 1700°C are required to vitrify the waste.
Another type of in situ soil decontamination process similar to vitrification is radio frequency (RF) in situ heating
(Dev et al., 1984). The process was originally developed to heat large amounts of earth, such as oil shale or tar
sand, to recover hydrocarbons. The waste site is heated with electromagnetic waves with frequencies between 2
and 45 MHz. A series of horizontal conductors are placed above the surface of the site and excited with a radio
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frequency generator. The organic pollutants are decontaminated by thermal decomposition, vaporization, and
distillation when heated for one to two weeks at temperatures between 300 to 400°C. The resulting gases that
move to the surface are collected by a vapor barrier and disposed of appropriately. The disadvantage of the RF
heating process is the large energy requirement. In addition, a large amount of the applied heat may be lost
because of thermal conduction. However, treatment of a hazardous waste landfill by RF in situ heating is thought
to be two to four times cheaper than incineration. The RF method is also advantageous because treatment is in
place, the waste is safely contained, small amounts of additional wastes are generated, and exposure to operating
personnel is minimal.
Another type of innovative physical process for in situ decontamination is the application of a vacuum in the
vadose (unsaturated) zone. The process can be used to monitor or recover subsurface contaminants from leaking
storage facilities, waste sites, or pipelines (Malot, 1985). Monitoring the unsaturated zone can provide
information concerning the potential for contamination of the ground water. Volatile organic compounds,
solvents, and petroleum-derived compounds are most amenable to treatment by this technique. Volatile tracers
can be added to tanks that contain nonvolatile compounds to verify potential leaks. The system is composed of a
subsurface extraction well and a vacuum source; recovery operations also include surface equipment for collection
or treatment and a probe submerged in the extraction well to monitor effectiveness. Each system design is site
specific and must be installed to accommodate the field constraints. The effectiveness of the technique has been
demonstrated (Agrelot et al., 1985). Vacuum extraction of organic contaminants is low cost because of the high
removal efficiencies in comparison to other techniques; the process can simultaneously remove free product,
residual hydrocarbons and vapors (Malot and Wood, 1985). In addition, decontaminating the vadose zone is
important because the source for pollutant entry or recharge into the aquifer is eliminated. Vacuum extraction can
also be used to recover floating product from perched water tables when recovery is difficult using a pumping
well; movement of floating product to an extraction well is difficult to induce when the perched water is thin.
The vacuum concept has also been used to solely vent vapors that result from hydrocarbon spills in the subsurface
(Crow and Minugh, 1985). The process was tested in a pilot-scale field study in which two test systems were
installed at a spill site where the hydrocarbon resided on shallow ground water; pure product was recovered using
conventional techniques. Each unit was composed of vapor monitoring probes, a vapor extraction well that
operated by negative pressure, and air inlet wells for entry of atmospheric air into the subsurface; two liquid ring
pumps supplied the vacuum source for venting. The results from the study indicated that the technique was
effective in controlling and eliminating vapors from the vadose zone; however, pulse venting may be more cost
effective because removal of the vapors required less time than reestablishing baseline vapor concentrations in the
vadose zone. The major limitation of the process is that vapors that are removed must be treated or released into
the atmosphere. The technique is largely undemonstrated and future studies are planned for further modification.
Coia et al. (1985) used positive pressure, or air stripping, to remove volatile chlorinated organics from
contaminated soils in a pilot field study. The volatile compounds were extracted from the vadose zone as a vapor
by forced ventilation. The design included a pipe vent, air ventillation system, and controls to monitor the
removal process. The stripping was accomplished by continuous injection and extraction of air through the pipe
vents; the extracted air is then passed through activated carbon before release to the atmosphere. The system was
effective in removing the chlorinated compounds from the vadose zone; however, further research is needed to
determine the effect of soil type and the presence of perched water on the process.
E. In Situ Biological Treatment
1. Microbial Activity In Aquifers
Microbial processes may be used to degrade contaminants in situ by stimulating the native microbial population.
Another in situ biostimulation technique which is not yet demonstrated is the inoculation of the subsurface with a
microbial population that has specialized metabolic capabilities. Even in the presence of an indigeneous
population which is acclimated to the organic contaminants, degradation may be limited at high contaminant
concentrations or by some environmental factor. Addition of electron acceptors, such as oxygen, and inorganic
nutrients, typically nitrogen, phosphorus, and trace metals, may provide the microflora with essential nutrients that
24
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are limiting in the presence of high concentrations of pollutants. Inoculation of a specialized microbial population
may reduce the time required for acclimation to the contaminants and/or allow the removal of recalcitrant
contaminants. Related processes such as the addition of bioemulsifiers or surfactants to increase the availability
of subsurface contaminants to the microflora can also be used. When applicable, biological processes may offer
the advantage of partial or complete destruction of the contaminants rather than simply transferring the pollution
to another phase of the environment.
Technologies for biorestoration of polluted aquifers have resulted from recent research indicating that subsurface
microorganisms exist, are metabolically active and often nutritionally diverse. A review, published by Dunlap
and McNabb (1973) of the Robert S. Kerr Environmental Research Laboratory, addressed subsurface biological
activity in relation to ground water pollution and initiated most of the research in this area. Before publication of
the review, the concept of biological activity below the rhizosphere had not been widely received.
Microbiologists were skeptical about biological activity in the subsurface because of oligotrophic conditions
below the rhizosphere (Leenheer et al., 1974) and an early study which had indicated that microbial numbers
decreased precipitously with depth (Waksman, 1916).
Sampling Methods for Subsurface Microbes-
A document that described sampling methods for subsurface microorganisms was published in 1977, by the
Environmental Protection Agency (Dunlap et al., 1977). The method for procuring a representative sample of
unconsolidated subsurface soil has since been modified (Wilson et al., 1983). A soil sample is collected by first
drilling a borehole to a desired depth with an auger and then taking the sample with a core barrel. After sample
procurement, the core is extruded through a sterile paring device that removes the outer layer of soil that has come
in contact with the core barrel. The remaining soil core is thus uncontaminated by the sampling procedure and
representative of the subsurface.
Investigations of microbial activity in the subsurface conducted prior to the development of the sampling
techniques were equivocal because of the potential for contamination during sample procurement. In addition,
many of the investigations were conducted using well water instead of core material. Recent evidence suggests
that the majority of subsurface microorganisms are associated with soil particles (Harvey et al., 1984). In
addition, well water may contain microorganisms that are artifacts of the well because of subsurface
contamination during well installation and changes in water quality around the well.
Microbial Numbers in the Subsurface--
Methods to enumerate the subsurface microflora also have been developed. Electron microscopy, viable counts,
epifluorescence microscopy, and measurements of biochemical components have been used to estimate microbial
biomass (Ghiorse and Balkwill, 1985; Ghiorse and Balkwill, 1983; Wilson et al., 1983; Smith et al., 1986;
Stetzenbach et al., 1986; Smith et al., 1985; Balkwill and Ghiorse, 1985; Bone and Balkwill, 1986; Webster et al.,
1985; White et al., 1983; Hoos and Schweisfurth, 1982; Ehrlich et al., 1983; Federle et al., 1986). In contrast to
Waksman's study (1916) which reported that microbial numbers declined with depth, uniform population levels
around 106-107 cells/g dry soil, measured by epifluorescence microscopy, were reported for profiles of
uncontaminated shallow aquifers (Ghiorse and Balkwill, 1985; Webster et al., 1985; Wilson et al., 1983; Ghiorse
and Balkwill, 1983; Balkwill and Ghiorse, 1985; Bone and Balkwill, 1986). However, bacteria in a chalk aquifer
(consolidated) were sporadically distributed with depth (Towler et al., 1985). Close examination of the subsurface
strata indicates patchiness of bacterial populations; samples from the top of the unsaturated zone of an artesian
aquifer yielded the highest counts whereas those from bedrock and confining layers yielded the lowest total counts
(Beloinetal., 1986).
Microbial Ecology of the Subsurface-
Bacteria are the predominant form of microorganism observed in the subsurface although a few higher life forms
have been detected (Wilson et al., 1983; Ghiorse and Balkwill, 1985; White et al., 1983). Some eucaryotic forms
25
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which may be fungal spores or yeast cells have been observed in the upper 10 m of a soil profile (Ghiorse and
Balkwill, 1983; Hoos and Schweisfurth, 1982; Federle et al., 1986). Bacteria, protozoa, and fungi have been
detected in samples of ground water collected from one-year-old wells (Hirsch and Rades-Rohkohl, 1983). In
addition, a slow-growing amoeba has been isolated and cultured from the ground water interface of an
uncontaminated soil (Balkwill and Ghiorse, 1985; Beloin et al., 1986).
Metabolic Activity of the Subsurface Microbial Community--
Organic matter that enters the uncontaminated subsurface is usually the more refractory humic substances which
resist degradation while percolating through the biologically active soil zone. The organic material available for
metabolism by the subsurface microflora is likely to be in low concentration and difficult to degrade. The
majority of microorganisms present in such nutrient-poor environments are generally oligotrophic.
Characterization of the subsurface microflora indicates that the bacteria are usually smaller (<1.0 mm in size) than
those in eutrophic environments and both Gram positive and negative cell types are present (Ghiorse and Balkwill,
1983; Wilson et al., 1983; Ghiorse and Balkwill 1985). Gram positive forms predominate in many
uncontaminated soils. The predominance of small, coccoid cells and hence a large surface to volume ratio for
enhanced nutrient uptake, is a likely mechanism for survival in an oligotrophic environment such as the
uncontaminated subsurface (Wilson et al., 1983). In contrast, subsurface soil contaminated with creosote waste
was found to contain more biomass and a greater proportion of Gram negative to Gram positive microbes when
compared to uncontaminated soil from the same site (Smith et al., 1985; Smith et al., 1986).
Studies have also indicated that many subsurface microorganisms are metabolically active. Of the total cell count,
about 0.01 to 50 percent can be recovered by plating on solid media and about 1 to 10 percent exhibit respiratory
activity measured by the reduction of 2-(p-iodophenyl)-3-p-nitrophenyl)-5-phenyl tetrazolium chloride by
cytochromes (Balkwill and Ghiorse, 1985; Webster et al., 1985). Microbial activity, measured by the hydrolysis
of fluorescein diacetate, declined with depth in the unsaturated zone of Ultisols and Alfisols (Federle et al., 1986);
however, 2-(p-iodophenyl)-3-(p-nitrophenyl)-5-phenyl tetrazolium chloride reduction varied greatly between
strata of a soil profile obtained from a shallow aquifer (Beloin et al., 1986).
Many subsurface microorganisms are nutritionally diverse (Table 2-3). Simple substrates such as glucose,
glutamic acid, arginine, a mixture of amino acids, and a synthetic compound, nitrilotriacetic acid, were
mineralized in samples of uncontaminated ground water (Larson and Ventullo, 1983). Polar solvents such as
acetone, isopropanol, methanol, ethanol, and tert-butanol also have been reported to degrade aerobically by
subsurface microorganisms (Novak et al., 1984; Jhaveri and Mazzacca, 1983). More challenging contaminants
that are aerobically degraded by subsurface microorganisms include the methylated benzenes, chlorinated
benzenes (Kuhn et al., 1985), chlorinated phenols (Suflita and Miller, 1985), and methylene chloride (Jhaveri and
Mazzacca, 1983). Highly lipophilic compounds such as naphthalene, methylnaphthalenes, dibenzofuran, fluorene,
and phenanthrene are also biotransformed in the subsurface (Wilson et al., 1985; Lee and Ward, 1985).
The microflora in some uncontaminated soils require little or no acclimation period to degrade many xenobiotics.
For example, toluene, chlorobenzene, and bromodichloromethane were biotransformed in uncontaminated soil,
but not 1,2 dichloroethane, 1,1,2-trichloroethane, trichloroethylene, and tetrachloroethylene (Wilson et al., 1983).
Benzene, toluene and the xylene isomers were found to degrade in uncontaminated subsurface soils (Barker and
Patrick, 1986). In addition, methanol (80-100 ppm) was degraded completely after two months, whereas tert-
butanol degraded much slower in two uncontaminated anaerobic aquifers (White et al., 1986).
In contrast to reports of degradation of xenobiotics in uncontaminated soil, long periods of acclimation to
subsurface pollutants may be required before biodegradation can occur. Wilson et al. (1985) reported degradation
of naphthalene, 1-methyl naphthalene, 2-methyl naphthalene, dibenzofuran and fluorene at 100-1000 mg/1 in
subsurface soil in the plume of contamination from a creosote waste pit; however, degradation of these
compounds was not observed in uncontaminated soil from the same site. The time and concentration required for
acclimation of the microflora to subsurface pollutants are unknown. Spain and Van Veld (1983) reported a
threshold concentration of 10 ppb for adaptation to p-nitrophenol in samples of sediment and natural water. A
better understanding of acclimation processes may explain why some chemicals persist in the subsurface even
though they have been reported to degrade in laboratory cultures and samples of water and soil.
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Table 2-3. Organic Compounds That Have Been Shown to be Biodegradable in the Subsurface
Soil from
Contaminated
Compound Area
Natural Compounds
glucose
glutamic acid
arginine
Solvents
acetone
ethanol
isopropanol
tert-butanol
methanol
bromodichloromethane
Aromatics
benzene
xylene
methylated benzenes
chlorinated benzenes
chlorinated phenols
naphthalene
dibenzofuran
fluorene
phenanthrene
toluene
chlorobenzene
no
yes
yes
no
no
yes
yes
yes
no
Aerobic
yes
yes
yes
yes
yes
yes
yes
yes
yes
Reference
Larson and Ventullo,
1983
Jhaveri and
Mazzacca, 1983
Novak et al., 1984
Wilson et al., 1983
Barker and Patrick,
1983
Kuhnetal., 1985
Suflita and Miller,
1985
Wilson etal., 1985;
Lee and Ward, 1985
Wilson et al., 1983
Environmental Factors Which May Limit Biodegradation-
Environmental factors may limit or preclude the biodegradation of subsurface organic pollutants, even in the
presence of adapted organisms. Recalcitrance of compounds thought to be biodegradable may result from lack of
an essential nutrient, substrate concentration, substrate inaccesibility.and the presence of toxicants (Alexander,
1975). Transport of contaminants in the subsurface also affects biodegradation. Transport is discussed in detail in
Section II.F.
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Biodegradation of many organic pollutants in the subsurface may be limited by insufficient oxygen. Alexander
(1980) reported that even the metabolism of carbohydrates may be inhibited in oxygen-depleted environments.
Lee and Ward (1985) found that the rate and extent of biotransformation of naphthalene, 2-methyl naphthalene,
dibenzofuran, fluorene, and phenanthrene were greater in oxygenated ground water than in oxygen-depleted
water. Contrary to general theory that complete degradation (mineralization) of hydrocarbons requires molecular
oxygen, more recent research suggests that alternate pathways exist under anaerobic conditions. Kuhn et al.
(1985) reported mineralization of xylenes in samples of river alluvium under denitrifying conditions. In addition,
benzene, toluene, the xylenes, and other alkylbenzenes were metabolized in methanogenic river alluvium that had
been contaminated with landfill leachate (Wilson and Rees, 1985); mineralization of toluene was confirmed by
adding 14C-labelled toluene and measuring the amount of 14CO2 produced. Grbic-Galic and Vogel (1986) also
reported mineralization of toluene and benzene under anaerobic conditions by a methanogenic consortium
acclimated to ferulate. Further tests indicated that water supplied the oxygen that is first incorporated into the
toluene and benzene ring (Vogel and Grbic-Galic, 1986).
The presence of oxygen may inhibit the biodegradation of many halogenated aliphatic compounds in the
subsurface. Degradation of trihalomethanes, trichloroethylene, and tetrachloroethylene did not occur in aerobic
cultures of sewage bacteria; however, the trihalomethanes were degraded anaerobically by mixed cultures of
methanogens (Bouwer et al., 1981). In addition, Bouwer and McCarty (1983b) reported that chloroform, carbon
tetrachloride and brominated trihalomethanes, but not chlorinated benzenes, ethylbenzene, or naphthalene were
biotransformed under denitrifying conditions.
In addition to oxygen, other nutrients may limit the biodegradation of organic pollutants in the subsurface.
Inorganic nutrients, such as nitrogen and phosphorous, may be limiting when the ratios of carbon to nitrogen or
phosphorous exceed that necessary for microbial processes. On the other hand, the presence of sulfate may inhibit
methanogenic consortia that have been reported to dehalogenate and mineralize many chlorinated aromatic
compounds (Suflita and Gibson, 1985; Suflita and Miller, 1985).
The effect of substrate concentration on biodegradation of organic compounds in surface soils and waters has been
documented (Alexander, 1985). Thresholds below which degradation is slow or does not occur may exist for
compounds that are readily biodegradable at higher concentrations. Boethling and Alexander (1979) reported that
less than 10 percent of 2,4-dichlorophenoxyacetate at concentrations of 22 pg/ml and 2.2 ng/ml was mineralized
in stream water whereas about 80 percent was mineralized at higher concentrations of 0.22 and 22 mg/ml. On the
other hand, microorganisms may be inhibited or killed by high concentrations of organic pollutants that result
from injection wells and hazardous waste sites. Lee (1986) reported that glucose mineralization was inhibited in
subsurface soil heavily contaminated with creosote; however, glucose was mineralized in uncontaminated and
slightly contaminated core material from the same site.
Other factors such as sorption, pH and temperature may also affect biodegradation of pollutants in the subsurface.
Many of the organic compounds contaminating the subsurface are highly lipophilic. These compounds are sorbed
by soil more strongly than the more hydrophilic compounds (Hutchins et al., 1985). Sorption may enhance
degradation by concentrating nutrients or conversely, prevent degradation by rendering the substrate unavailable
to the microorganism. Zobell (1943) reported that sorption of organic material to solid surfaces in dilute nutrient
solutions increased microbial respiration. In contrast, Ogram et al. (1985) observed that 2-4 dichlorophenoxy
acetic acid sorbed to soil was completely protected from microbial degradation. Therefore, sorption may be
important in nutrient scavenging in uncontaminated aquifers which are generally oligotrophic; however, sorption
may compete with the microflora for subsurface pollutants that are relatively hydrophobic.
The soil pH may affect sorption of ionizable compounds in addition to limiting the types of microorganisms in the
subsurface. Methanogens, which have been implicated in mineralization of some aromatic hydrocarbons, are
inhibited at pH values less than 6.0 (Alexander, 1977). Nitrification, the microbial conversion of ammonia to
nitrate, is also limited at pH values below 6.0 and is negligible below 5.0. Hambrick et al. (1980) also reported
that mineralization of octadecane and naphthalene in sediment was faster at a pH of 8.0 than 5.0.
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Temperature also influences microbial metabolism of subsurface pollutants. The temperature of the upper 10 m
of the subsurface may vary seasonably; however, that between 9-18 m approximates the mean air temperature
(between 3 and 25°C in the United States) of a particular region (McNabb and Dunlap, 1975). Biodegradation of
subsurface pollutants in the more northern climates may therefore be limited by cooler temperatures.
Bartholomew and Pfaender (1983) reported that the microbial metabolism of m-cresol, nitrilotriacetic acid, and
chlorinated benzenes in fresh water and estuarine areas decreased as temperature decreased. Atlas (1975) and
Mulkins-Phillips and Stewart (1974b) also reported a direct relationship between petroleum hydrocarbon
degradation and temperature.
In summary, the subsurface environment contains microbes that degrade many of the organic compounds that
contaminate ground water. The subsurface microorganisms in uncontaminated aquifers are likely to be
oligotrophic. The majority of the microorganisms are associated with soil particles. Even in the presence of
adapted populations, environmental factors such as temperature, pH, dissolved oxygen levels, inorganic nutrient
concentrations, and the availability and concentration of the organic contaminants may limit biodegradation of
subsurface pollutants.
2. Biostimulation by Addition of Limiting Nutrients
Development of the In Situ Biostimulation Process with Oxygen Supplied by Air Sparging
Application of the degradative activity of subsurface microbes-The potential for biodegradation of organic
compounds in contaminated aquifers was first reported in 1971. Bacteria capable of degrading hydrocarbons were
observed in an area contaminated with gasoline; however, biodegradation of the gasoline was limited by the
availability of oxygen, mineral nutrients, and hydrocarbon surface area (Williams and Wilder, 1971). Williams
and Wilder (1971) suggested that these hydrocarbon-degrading bacteria could be used to clean the aquifer of
residual gasoline; however, concern was expressed that bacterial growth would plug the well and formation.
Davis et al. (1972) recommended supplying the indigenous microflora with nutrients, oxygen, and moisture rather
than inoculating the subsurface with commerical biological products such as dried bacterial cultures. Oxygen-
limited degradation of hydrocarbons was reported by McKee et al. (1972) in studies designed to investigate the
fate of gasoline trapped in the pore space of sand columns. Several species of Pseudomonas and Arthrobacter
were isolated from ground waters associated with a gasoline spill and used in the column experiments. The total
number of gasoline-degrading bacteria in the ground water numbered over 50,000 cells/ml in the contaminated
zone, but less than 200 cells/ml had been found in the uncontaminated wells and in wells where gasoline had not
been detected for a year. The presence of high numbers of gasoline-degrading bacteria was suggested as an
indicator of cleanup progress. In the column study, the bacteria rapidly degraded the gasoline in the zone of
aeration but slowly degraded that in the saturated zone. In a similar study, Litchfield and Clark (1973)
enumerated hydrocarbon-degrading bacteria in ground waters from 12 sites which were contaminated with
petroleum. The numbers of hydrocarbon-degrading bacteria ranged from 103 to 106 cells/ml, with similar numbers
of both aerobic and microaerophilic organisms, in ground waters containing more than lOppm hydrocarbon.
Hydrocarbon-degrading bacteria were found in ground water from all 12 sites; however, on a site by site basis,
there were no relationships between the types of organisms, the type of petroleum contamination, the geological
characteristics, or the geographical location of the site.
Application of the degradative capacity of subsurface microorganisms to restore gasoline-contaminated ground
water was first demonstrated by Raymond, Jamison, Hudson and coworkers at Suntech (Lee and Ward, 1985). In
1974, Raymond (1974) received a patent on a process designed to remove hydrocarbon contaminants from ground
waters by stimulating the indigenous microbial population with nutrients and oxygen. Oxygen and nutrients are
introduced into the formation through injection wells and production wells were used to circulate them through
the aquifer. Placement of the wells was dependent on the area of contamination and the porosity of the formation,
but usually no closer than 100 ft apart. The nutrient amendment consists of nitrogen, phosphorus, and other
inorganic salts, as required, at concentrations of 0.005 to 0.02 percent by weight; oxygen was supplied by sparging
air into the ground water. The process was projected to require about six months to achieve degradation of 90
percent of the hydrocarbons if the growth rate of the microorganisms was 0.02 g/L per day. The numbers of
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bacterial cells were expected to return to ambient levels once the addition of nutrients was terminated. The
process was expected to be more efficient in treating ground water contaminated with less than 40 ppm of
gasoline.
First application of the biostimulation process-A pipe line leak in Ambler, Pennsylvania was the first site where
Raymond's patent on biorestoration was demonstrated. An estimated 380,000 L of high octane gasoline had
leaked into a highly fractured dolomite outcrop underlaid by quartzite (Raymond et al., 1975). Depth to the water
table ranged from 9.2 to 30.5 m in the 46 monitoring wells installed at the site. Before biorestoration was
attempted, conventional pump and treat technologies were used as remedial action. Containment of the gasoline
was achieved by continuously pumping water from wells located in the spill area. About 238,000 L of the
gasoline was recovered by physical methods; however, the recovery program was incomplete and approximately
119,000 L of residual gasoline remained. The concentration of dissolved gasoline in the withdrawn ground water
averaged less than 5 ppm. The time required for restoration of the aquifer using this pump and treat technique
was estimated to be more than 100 years.
Problems in analyzing the concentration of residual hydrocarbons during the pump and treat phase were later
attributed to the presence of hydrocarbon-degrading bacteria (Raymond et al., 1975). A program designed to
investigate the potential for biodegradation of the gasoline by these organisms was then initiated. A laboratory
study indicated that supplements of air, inorganic nitrogen, and phosphate salts could increase the numbers of
hydrocarbon-degrading bacteria by one thousand-fold (Raymond et al., 1976). Small scale field studies also
indicated that nutrient additions would enhance the growth of bacteria that degrade hydrocarbons (Jamison et al.,
1975). A full scale program to stimulate the biodegradation of the gasoline in the aquifer was then initiated
(Raymond et al., 1976). The nutrient amendment, which contained ammonium sulfate, disodium phosphate, and
monosodium phosphate, was injected into the aquifer as a 30 percent concentrate by batch addition. Either
ammonium or nitrate could serve as the nitrogen source. Magnesium, calcium, and iron were not included in the
concentrate because the small scale field study indicted that these inorganic nutrients were not limiting (Jamison
et al., 1975). Biodegradation of 1 liter of gasoline was estimated to require 44 g of nitrogen, 22 g of phosphorus,
and 730 g of oxygen. However, Baehr and Corpcioglu (1985) estimated that degradation of a pound (0.63 liter) of
gasoline requires 3.5 g of oxygen. Batch addition of the nutrients worked as well as continuous addition and was
more cost-effective; however, high concentrations of nutrients may osmotically shock the microorganisms
(Raymond et al., 1976). Oxygen was supplied by sparging air into the wells using paint sprayer-type compressors
and Carboundum diffusers with a flow rate of 0.06 m3/min. As a result, the bacterial population increased from
about 103 to 107 cells/ml. High bacterial counts mirrored locations of high gasoline concentrations at the site
(Raymondetal., 1975).
During the biostimulation program at the Ambler, Pennsylvania site, 32 cultures of bacteria that actively
metabolized gasoline were isolated and characterized; the isolates included species of the genera Nocardia.
Micrococcus. Acinetobacter. Flavobacterium. and Pseudomonas: some cultures could not be identified. Studies
were conducted to determine the metabolic capabilities of these isolates (Jamison et al., 1976). The data
suggested that the Nocardia cultures were largely responsible for the degradation of the aliphatic hydrocarbons
whereas those from the genus, Pseudomonas. degraded the aromatics. Branched paraffins, olefins, or cyclic
alkanes did not support the growth of any isolate. Co-oxidation may have played a major role in the
biodegradation of these organics. An alternative hypothesis is that the bacteria capable of degrading these
compounds were not isolated. The lack of microbial growth on some types of hydrocarbons may result from the
toxicity or structure of the substrate. Straight chain aliphatics which are less than 10 carbons in length can be
toxic whereas longer chains and branched alkanes are often resistant to microbial attack (Suflita, 1985).
Substitutions on aromatics that are biodegradable may render them recalcitrant. Huddleston et al. (1986) gave the
following order for petroleum hydrocarbon constituents, in order of decreasing biodegradability: linear alkanes
C1019, gases C24, alkenes C5 9, branched alkenes C12, alkenes C3 u, branched alkenes, aromatics, and cycloalkanes.
The bioreclamation program conducted by Suntech at Ambler, Pennsylvania,was reasonably successful. During
the period of nutrient addition, the concentration of gasoline in the ground water did not decline; however gasoline
could not be detected in ground water 10 months later (Raymond et al., 1976). A thousand-fold increase in the
numbers of total and hydrocarbon-degrading bacteria was observed in ground water from many wells (Raymond et
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al., 1975). The waters from some wells exhibited foaming because of high microbial numbers and associated
exopolysaccharides. Counts of microorganisms determined one year after nutrient addition was terminated
indicated that the microbial population had declined. Estimates based on the amount of nitrogen and phosphorus
removed from the nutrient solution suggested that between 88,600 and 112,400 L of gasoline were degraded.
However, this estimate was not particularly accurate because some of the nutrients may have been adsorbed by
soil or lost from the biostimulation area by dilution. In addition, the estimates were based on discrete samples
rather than composited samples. Large quantities of nutrients were used in this project; approximately 79 metric
tons of food grade reagents were purchased.
Steps in the biostimulation process-The Ambler, Pennsylvania site case history is an example of the
biostimulation process. The basic steps involved in an in situ biorestoration program are the following: 1) site
investigation; 2) free product recovery; 3) microbial degradation enhancement study; 4) system design; 5)
operation; and 6) monitoring (Lee and Ward, 1986). The first step in the process is to define the hydrogeology
and the extent of contamination of the site. Important hydrogeologic characteristics include the direction and rate
of ground water flow, the depths to the water table and to the contaminated zone, the specific yield of the aquifer,
and the heterogeneity of the soil. In addition, hydraulic connections between aquifers, potential recharge and
discharge areas, and fluctuations in the water table must be considered. The sustainable pumping rate must also
be determined (Roux, 1985; Brown et al., 1985a). These parameters can be determined by surveying the existing
data for that site and region, reconnaissance by experienced hydrogeologists, geophysical surveys, excavation of
test pits, and installation of boreholes and monitoring wells (Josephson, 1983). Low dissolved oxygen
concentrations may indicate an active zone of hydrocarbon biodegradation (Chaffee and Weimer, 1983). The
types and concentrations of contaminants is also important (Brown et al., 1985a). The type of remedial action
chosen depends on the time elapsed since the spill, the area! extent of contamination, the nature of contaminants
and whether the contamination is acute, chronic, or periodic. The urgency for action and the treatment level that
must be achieved will depend on the potential for contamination of drinking water or agricultural water wells.
After defining the site hydrogeology, the next step is recovery of free product. Depending on the characteristics
of the aquifer and contaminants, free product can account for as much as 91 percent of the spilled hydrocarbon
(Brown et al., 1985a). The remaining hydrocarbon, which is sorbed to the soil and dissolved in the ground water,
may account for 9 to 40 percent of the total hydrocarbon spilled; the majority is usually sorbed, however, the
dissolved phase is the most difficult to treat. The pure product can be removed using techniques described in
sections II B.2. an D. Physical recovery often accounts for only 30 to 60 percent of the spilled hydrocarbon before
yields decline (Yaniga and Mulry, 1985).
Prior to in situ treatment, a laboratory study is conducted to determine the nutrient requirements that will enable
the indigenous microorganisms to efficiently degrade the contaminants (Lee and Ward, 1985b). Kaufman (1986)
suggested that these laboratory studies can provide a reliable basis for field trials; however, the studies must be
performed under conditions that simulate the field. For example, Kuhlmeier and Sunderland (1986) conducted a
laboratory investigation of the unsaturated zone using samples saturated with ground water. Clearly, the results of
their study do not represent the fate of the organics in the unsaturated zone. A chemical analysis of the ground
water provides little information about the nutrient requirements of the microflora (Raymond et al., 1978).
However, the chemistry of the site will affect the nutrient formulation. For example, large quantities of oxygen
may be consumed to oxidize reduced iron (Hallberg and Martinell, 1976). In addition, nutrients may sorb onto
soils, especially silts and clays and be unavailable to the microflora (Brubaker and Crockett, 1986). Limestone
and high mineral content soils and ground waters will also affect nutrient availability by reacting with the
phosphorus.
Nutrient requirements are usually site specific. Nitrogen and phosphorus were required at the Ambler site
(Raymond et al., 1976a); however, the addition of ammonium sulfate, mono-and disodium phosphate, magnesium
sulfate, sodium carbonate, calcium chloride, manganese sulfate, and ferrous sulfate was required at other sites
(Raymond et al., 1978; Minugh et al., 1983). The form of the nutrient may also be important; ammonium nitrate
was less efficient than ammonium sulfate in one aquifer system.
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Laboratory studies conducted to determine appropriate nutrient formulations can be performed using a number of
techniques. An increase in the number of total and hydrocarbon degrading bacteria has been used to identify
limiting nutrients in a factorial experimental design (Raymond et al., 1976,1978). However, an increase in
microbial numbers does not demonstrate that the substrate of interest is being used. Batch culture techniques
designed to measure the disappearance of the contaminant (Flathman and Githens, 1985) and electrolytic
respirometer studies designed to measure the uptake of oxygen also have been used (Flathman et al., 1985). The
results of another laboratory investigation indicated that dissolved oxygen was the primary factor limiting
biodegradation of aromatic contaminants at a wood creosoting site rather than inorganic nutrients (Lee, 1986).
Biotransformation studies which measure the disappearance of the contaminants or mineralization studies which
indicate the complete destruction of the compound to carbon dioxide and water will confirm that the contaminants
are being degraded. Controls to detect abiotic transformation of the pollutants and tests to detect toxic effects of
the contaminants on the microflora should be included (Flathman et al., 1984).
A system for injection of nutrients into the formation and circulation through the contaminated portion of the
aquifer must be designed and constructed (Lee and Ward, 1985b). The system usually includes injection and
production wells and equipment for the addition and mixing of the nutrient solution (Raymond, 1978). A typical
system is shown in Figure 2-1. Placement of injection and production wells may be restricted by the presence of
physical structures. Wells should be screened to accommodate seasonal fluctuations in the level of the water
table. Air can be supplied with carborundum diffusers (Raymond et al., 1975), by smaller diffusers constructed
from a short piece of DuPont Viaflo tubing (Raymond et al., 1978), or by diffusers spaced along air lines buried in
the injection lines (Minugh et al., 1983). The size of the compressor and the number of diffusers are determined
by the extent of contamination and the time allowed for treatment (Raymond, 1978). Nutrients also can be
circulated using an infiltration gallery (Figure 2-2); this method provides an additional advantage of treating the
residual gasoline that may be trapped in the pore spaces of the unsaturated zone (Brenoel and Brown, 1985).
Oxygen also can be supplied using hydrogen peroxide, ozone, or soil venting (see section on alternative oxygen
sources). Well installation should be performed under the direction of a hydrogeologist to ensure adequate
circulation of the ground water (Lee and Ward, 1985b). Produced water can be recycled to recirculate unused
nutrients, avoid disposal of potentially contaminated ground water, and avoid the need for makeup water.
Inorganic nutrients can be added to the subsurface once the system is constructed. Continuous injection of the
nutrient solution is labor intensive but provides a more constant nutrient supply than a discontinuous process.
Continuous addition of oxygen is recommended because the oxygen is likely to be a limiting factor in
hydrocarbon degradation.
The performance of the system and proper distribution of the nutrients can be monitored by measuring the
organic, inorganic, and bacterial levels (Lee and Ward, 1985b). Carbon, dioxide levels are also an indicator of
microbial activity in the formation (Jhaveri and Mazzacca, 1985). Depending on the charcteristics of the nutrients
and soil, nutrients can be removed from solution by sorption onto soil (Brubaker and Crockett, 1986). About 90
percent of the ammonium and phosphate and 70 percent of the hydrogen peroxide added to a sandy soil with low
calcium, magnesium, and iron was recovered. After passage of a nutrient solution through a column packed with
a clay soil that had high calcium and magnesium but low iron and chloride levels, 100,66 and 25 percent of the
ammonium, phosphate, and hydrogen peroxide were recovered, respectively. However, after passage of a nutrient
solution through a column packed with a clay soil high in calcium, magnesium, and chloride, but low in iron, 75,
100, and 15 percent of the ammonium, phosphate, and hydrogen peroxide, respectively, were recovered. Both soil
and ground water samples should be collected and analyzed to fully evaluate the treatment effectiveness (Roux,
1985). Raymond et al. (1975) reported that the most difficult problem in optimizing microbial growth in the
Ambler reservoir was the distribution of nutrients, which was made difficult by the heterogeneity of the dolomite
formation.
Additional case histories in which oxygen was supplied by air sparging-In situ biorestoration has been
largely used to treat gasoline spills and with reasonably good success. However, many of the reports on in situ
biorestoration lack sufficient data to fully judge the overall effectiveness and costs associated with the process.
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To Sewer or
Recirculate
Air Compressor
• Production Well
-Xr
V J
Nutrient
Addition
Tank
Coarse Sand
Water Table
-Xh
Water Supply
-Injection Well
Figure 2-1. Typical schematic for aerobic subsurface biorestoration.
Air Compressor or
Hydrogen Peroxide
Tank
Nutrient Addition
V
Monitoring Well
Water Table
Recovery Well
Figure 2-2. Use of infiltration gallery for recirculation of water and nutrients in in situ biorestoration.
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In a high permeability sand aquifer contaminated with hydrocarbons in Millville, New Jersey, the in situ
biorestoration program was successful in removing free product, but residual hydrocarbon was found at the last
sampling period (Raymond et al., 1978). The nutrient solution was moved through the formation at rates of 8 to
14 ft/day, but dissolved oxygen was rapidly consumed and did not increase in some of the main wells at all.
However, analysis of core material collected from the aquifer indicated that the concentration of gasoline had not
changed substantially during the biostimulation program. During the initial treatment process, inadequate
dissolved oxygen levels led to the microbial formation of phenol, but the phenol levels declined as more aerobic
conditions were achieved. A ten to one thousand-fold increase in the number of gasoline-utilizing bacteria was
noted in the area with the highest gasoline levels. The cleanup met the state requirement of removal of the free
gasoline and was subsequently stopped.
At a gasoline spill in La Grange, Oregon, nine months of treatment by in situ biorestoration and a vapor
elimination program succeeded in removing the free product and mitigating the vapor problems at two restaurants
(Minugh et al., 1983); however, the concentration of gasoline in the pits in the biorestoration treatment area still
ranged from 100 to 500 ppm in the majority of the samples. After an additional three months of treatment, the
dissolved organic levels in the ground water had decreased from an average of 20 ppm to less than 5 ppm in the
majority of the samples.
Fumes released from a pipeline spill of gasoline temporarily closed an elementary school (Suntech, 1978). A
pumping well was used to maintain the water table below the school's foundation and physical recovery was used
to remove two-thirds of the gasoline. An enhanced biodegradation program was initiated by circulating nutrients
and oxygen through the formation for six months. After the cleanup, hydrocarbons could not be detected and the
fumes that had threatened the school had been eliminated.
Minimum hydrocarbon concentrations achievable bv in situ biostimulation-The minimum concentration of
hydrocarbon that can be achieved by in situ biorestoration is unknown and is most likely site specific. A natural
gradient field test in a sandy Canadian aquifer required 434 days to reduce 1,000 to 2,400 ppb of benzene, toluene,
and the xylene isomers to below the detection limits (1 to 2 ppb) in the absence of added nutrients and oxygen
(Barker and Patrick, 1986). The distribution of dissolved oxygen in the plume was heterogeneous and probably
controlled biodegradation of the aromatics.
Jensen et al. (1986) suggested that the indigenous microflora should be able to reduce the concentration of
hydrocarbons below 1 mg/L when the initial hydrocarbon concentration is less than 10 mg/L and adequate
quantities of nutrients and oxygen are supplied. The results of batch experiments using ground water from
hydrocarbon-contaminated aquifers showed that the native microflora could generally reduce the concentrations of
toluene, benzene, xylene, trimethyl benzene, naphthalene, methyl naphthalene, biphenyl, ethyl naphthalene, and
dimethyl naphthalene from a range of 400 to 1,100 mg/L to less than 1 mg/L within a week in the presence of
oxygen and nutrients; however, phenanthrene and toluene persisted at higher concentrations in two of the ground
waters after incubation for six days.
The concentration of trace level organics in an aquifer may be reduced by providing a primary substrate that
supports microbial growth and allows the organisms to act upon the trace level organics as secondary substrates
(Bouwer, 1984). The concentration of the trace organic or secondary substrate is thought to be below the
minimum substrate concentration (S^) required to support microbial growth (Rittman and Kobayashi, 1982). The
S^ concept was developed to describe limitations related to transport of organics into a biofilm and the
subsequent kinetics of reaction. There are several examples of S^. A reactor fed laboratory grade water
containing 0.59 mg/L TOC was able to reduce acetate below the S^ value (0.03 mg/L) for acetate. Shimp and
Pfaender (1985) demonstrated that addition of fatty acids, carbohydrates and amino acids enhanced the ability of
mixed microbial populations to degrade substituted phenols. These data suggest that the addition of naturally
occurring substrates may enhance the biodegradation potential of some xenobiotics. However, the addition of a
primary substrate may not support the removal of some compounds. A biofilm supported by thymine could utilize
alanine and acetate, both common metabolites, but not phenol and galactose (Rittman and Kobayashi, 1982).
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Treatment trains--In many hydrogeologic systems which become contaminated from leaking underground storage
tanks, a remediation process may be so complex in terms of contaminant behavior and site characteristics that no
one system or unit will meet all requirements. Very often, it is necessary to combine several unit operations, in
series and sometimes in parallel, into one treatment process train in order to effectively restore ground water
quality to a required level (Wilson et al., 1986). Examples of treatment trains include:
(l)physical containment with product removal and surface treatment;
(2)product removal with unsaturated zone flushing followed by in siiu chemical treatment;
(3)physical containment with in situ physical/chemical treatment; and
(4)product removal followed by in situ biological treatment.
Physical containment through barriers and hydrodynamic controls alone merely act as temporary plume control
measures. However, hydrodynamic processes must also be integral parts of any withdrawal and treatment or in
situ treatment measures. Most remediation projects where enhanced biorestoration has been applied have started
by removing heavily contaminated soils. This was usually followed by installing pumping systems to remove free
product floating on the ground water, before biorestoration enhancement measures were initiated to degrade the
more diluted portions of the plume.
There are numerous proven surface treatment processes available for treating a variety of organic and inorganic
wastewaters. However, regardless of the source of ground water contamination and the remediation measures
anticipated, the limiting factor is getting the contaminated subsurface material to the treatment unit or units, or in
the case of in situ processes, getting the treatment process to the contaminated material. The key to success is a
thorough understanding of the hydrogeologic and geochemical characteristics of the area. Such an understanding
will permit full optimization of all possible remedial actions, maximum predictability of remediation
effectiveness, minimal remediation costs, and more reliable cost estimates (Wilson et al., 1986).
The role of biorestoration in combination treatment schemes is often difficult to assess. Yaniga et al. (1985a)
described the cleanup of a gasoline spill in which an air stripper was use4 to reduce the contaminants in the
withdrawn ground water and to supply oxygen before the water was recirculated to the aquifer via an infiltration
gallery. Before recirculation, ammonium chloride, sodium monophosphate, sodium diphosphate, iron sulfate, and
manganese sulfate were added in slug batches to the treated water. Additional oxygen was supplied by sparging
air into the wells. As a result, the dissolved oxygen increased from a range of 0-5 to 5-10 ppm; the hydrocarbon
degrading bacteria increased from 102-103 to lO'-lO4 cells/ml with just oxygen addition by air stripping and
sparging and then increased to 106 cells/ml with nutrient addition and additional oxygen. Brown et al. (1985b)
identified another gasoline contaminated aquifer which was treated using air sparging. An estimated 25,000 to
30,000 gallons of gasoline entered a 20 ft thick coarse grain sand and fine gravel aquifer. Recovery of free
product accounted for 18,500 gallons of the spilled gasoline; however, an estimated 10,000 gallons was sorbed to
the soil at concentrations of 2,000 to 3,000 ppm and 30 to 40 ppm was dissolved in the ground water. The
concentration of gasoline was reduced to less than 50 ppm in the soil and less than 1 ppm in the ground water by
air sparging. Only 1 to 2 ppm of dissolved oxygen could be achieved in the wells by air sparging.
A spill of four solvents-methylene chloride, n-butanol, acetone, and dimethyl aniline—into a glacial till aquifer
was withdrawn and treated by an activated sludge process, allowed to settle, and then recharged into the
subsurface through injection trenches after being aerated and amended with nutrients (Jhaveri and Mazzacca,
1983). The recharge water contained organisms acclimated to the solvents in addition to a nutrient amendment
containing nitrogen, phosphate, magnesium, sulfate, carbonate, manganese, and iron. Additional oxygen was
supplied to the aquifer using a series of injection wells. Removal efficiencies of methylene chloride, n-butanol,
and acetone were greater than 97 percent and the dimethyl aniline levels were reduced by greater than 93 percent
in the above ground treatment. The concentrations of the solvents in the resulting effluent decreased to 0.04 mg/L
for n-butanol, 0.92 mg/L for methylene chloride, 0.18 mg/L for dimethyl aniline, and 1.12 mg/L for acetone from
initial concentrations of 19.1,58.5,2.9, and 38.8 mg/L, respectively. Based upon COD and gas chromatography
analysis, the plume was reduced in size by 90 percent after three years of operation (Jhaveri and Mazzacca, 1985).
The COD was reduced from 300 to 20 mg/L in one monitoring well. Based on the rate of ground water flow, this
reduction in COD coincided with the expected arrival time of the treated ground water at that well. Elevated
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levels of carbon dioxide in ground water collected from the treatment zones, in comparison to those observed in
uncontaminated and decontaminated wells, suggested that in siiu biorestoration was occurring. However, the
solvents were detected in the ground water beyond the projected date for completion of the project and the New
Jersey Department of Environmental Protection standards had not been achieved after three years of operation.
Flathman et al. (1985) and Quince et al. (1985) discussed cleanup of a methylene chloride spill using physical and
biological above-ground treatment processes and in situ biological treatment. Following sand filtration to remove
particulates, air stripping, combined with a heat exchanger to improve stripping efficiency, was initially used to
treat the withdrawn ground water and the water was used to flush the soil (Quince et al., 1985). Air stripping
removed about 98 to 99.9 percent of the methylene chloride in the withdrawn water. The concentration of
methylene chloride in the ground water was reduced by 97 percent in one downstream monitoring well.
Biological treatment was used to further reduce the concentration of the methylene chloride after addition of
ammonia and phosphate. An activated sludge unit was seeded with acclimated organisms from a wastewater
treatment plant receiving methylene chloride and these organisms were used to inoculate the soil (Flathman et al.
1985). After 43 days of in situ biological treatment, the concentration of methylene chloride in ground water from
a monitoring well 20 ft from the spill declined from 192 to 6 ppm and 156 ppm chloride was released; however, it
could not be determined whether the added bacteria or indigenous microflora or both were involved in methylene
chloride degradation. Both air stripping and biological treatment removed 99.9 percent of the initial amount of
methylene chloride present during the four months of field operation. The concentration of methylene chloride
was reduced from 20,000 to less than 1 ppm in the source wells (Quince et al., 1985).
The subsurface at the Naval Air Engineering Center in Lakehurst, New Jersey, was contaminated with ethylene
glycol resulting from the loss of about 4,000 gallons of cooling water from a lined surface storage lagoon
(Flathman et al., 1984). The unsaturated zone was contaminated with concentrations of ethylene glycol as high as
4,900 mg/kg soil whereas ethylene glycol in the ground water was 2,100 mg/L. The highly contaminated soils
were treated using injection and recovery wells whereas the ground water contaminated with ethylene glycol was
treated by an above ground activated sludge unit and by adding ethylene glycol degrading bacteria and nutrients to
the subsurface (Flathman et al., 1985). A biofeasibility study using an electolytic respirometer had demonstrated
that the concentration of ethylene glycol could be reduced to less than 50 ppm within ten days by the natural
microflora in the ground water and that the concentration of ethylene glycol at 1,300 ppm was not toxic. The
initial operational phase was designed to degrade as much of the ethylene glycol as possible by treatment above
ground with an activated sludge unit. The effluent from the activated sludge unit was amended with oxygen,
nitrogen, and phosphorus, adjusted to neutral pH, and then reinjected into the subsurface to create a closed-loop
system. The amended effluent was used to flush the contaminated soil and inoculate the ground water with
nutrients and acclimated bacteria. The concentration of ethylene glycol in ground water collected from the plume
recovery wells was reduced from 420-690 ppm to less than 50 ppm within 26 days (Flathman and Caplan, 1985);
however, the unsaturated zone still contained pockets of ethylene glycol. A passive treatment system which
involved adding lime and diammonium phosphate to the soil surface continued after termination of the active
biorestoration phase. By the end of the treatment program, ethylene glycol could not be detected (detection limit,
50 ppm) in ground water collected from the production wells.
A shallow basin comprised of sand and pea gravel was contaminated with isopropanol and tetrahydrofuran
(Flathman and Githens, 1985). In addition to isopropanol and tetrahydrofuran, acetone was also detected in the
ground water and believed to be a byproduct of isopropanol degradation. Remedial action consisted of a recovery
system, an above ground biological reactor, and recharging the aquifer with the effluent from the reactor which
created a closed-loop system. The effluent, which contained acclimated bacteria, was also amended with nutrients
before reinjection into the subsurface. The soils were flushed with the treated ground water to remove sorbed
organics and introduce acclimated organisms into the aquifer. Maximum concentrations of isopropanol (950 ppm)
and acetone (190 ppm) were detected in ground water from a centrally located well as a result of flushing pockets
of contamination from the subsurface. The pattern of change in isopropanol and acetone concentrations was
similar. The concentration of acetone in the ground water increased initially until the majority of the isopropanol
had been degraded, and then declined to less than 0.2 ppm. Extrapolations from the data indicated that 99 percent
of the contaminants would be removed within 33 days. Estimated cost for removal and disposal of 200,000 ft3 of
contaminated soil was $550,000 whereas the biological treatment program was estimated to cost one-fifth as
much.
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Winegardner and Quince (1984) documented two additional case histories of in siiu biorestoration that involved
addition of acclimated bacteria. The first case history described the cleanup of a train derailment which released a
semi-soluble aliphatic hydrocarbon plasticizer. Recovery wells were used to collect the plasticizer from the
subsurface. Later, surface recharge and shallow injection were used to flush the plasticizer out of the soil; this
treatment reduced the peak concentration of greater than 2,000 ppm in a widespread area to a much smaller zone
after 70 days, in addition to reducing the concentration of the plasticizer throughout the contaminated area. Air
stripping and carbon adsorption were used initially; however, these techniques were replaced by biological
treatment using activated sludge. The water treated by activated sludge was used as an inoculant to introduce the
acclimated bacteria into the subsurface to enhance in situ biorestoration. The concentration of the plasticizer in
the recovered water was reduced from approximately 1,700 to 400 ppm after clarification; however, the
importance of each component in the treatment process could not be determined.
The second case history involved contamination of a glacial kame deposit of sand, gravel, silt, and clay with
chloroform from a leaking pipeline. Ground water was withdrawn and treated with a mixed media prefilter, an
activated sludge bioreactor and settling vessel, and a heated air stripper. The effluent from the activated sludge
bioreactor was used as an inoculant for biorestoration. The effluent from the air stripper was discharged into a
process sewer or into the subsurface. A forced flushing/recovery system was used to enhance the recovery of the
chloroform. Biological treatment followed the physical recovery; however, treatment effectiveness was not
discussed.
Alternate Oxygen Sources
The supply of dissolved oxygen may limit in situ, biorestoration of hydrocarbons, especially in low permeability
aquifers (Raymond et al., 1978). Depending upon the temperature of the ground water, only 8 to 12 mg/L of
dissolved oxygen can be achieved by air sparging, and incomplete transfer of oxygen into water may reduce this
even further. Using only the oxygen provided by sparging air into the ground water, 1,500 to 5,400 pore volumes
of air would be required to completely degrade the hydrocarbon in an aquifer that is six feet deep, one acre in size,
has a porosity of 30 percent, and contains 4000 mg/L gasoline (Brown et al., 1984).
Alternative sources of oxygen include pure oxygen, hydrogen peroxide, and ozone. Other methods of supplying
oxygen to the subsurface are soil venting or air flooding (Wilson and Ward, 1986) and colloidal dispersions of air
in a surfactant matrix (Michelsen et al., 1985). Concentrations of 40 to 50 mg/L of dissolved oxygen can be
achieved with pure oxygen; however pure oxygen is somewhat expensive, may bubble out of solution before the
microflora can use it, and is extremely flammable (Brown et al., 1984).
Hydrogen peroxide-Hydrogen peroxide, decomposes to form one molecule of water and half of a molecule of
oxygen, and can be used as a source of oxygen (Equation 2-1).
ff2O2 - H2O + 1/2 02 (2-1)
However, hydrogen peroxide is used as a sterilant at concentrations of 3 percent and levels as low as 200 ppm can
be toxic to microorganisms. Ground water organisms inoculated into sand columns could tolerate 0.05 percent
hydrogen peroxide, but higher levels were toxic (Britton and Texas Research Institute, Inc., 1985) In a study
designed to investigate the effect of increasing concentrations of hydrogen peroxide on gasoline degradation, the
culture acclimated to hydrogen peroxide levels that were gradually increased from 0.05 to 0.2 percent; however,
removal of gasoline was not greatly increased in comparison to the control without hydrogen peroxide. Microbial
counts were higher in columns in which hydrogen peroxide was incrementally increased than those which
received 0.05 percent hydrogen peroxide. These data suggest an oxygen limitation at lower concentrations of
hydrogen peroxide. Large populations of microorganisms survived high hydrogen peroxide concentrations better
than small populations (Texas Research Institute, 1982). In a column study in which oxygen concentration was
varied from 8 to 200 ppm using air, 60 percent nitrogen/40 percent oxygen, pure oxygen, or a hydrogen peroxide
solution, microbial growth and gasoline degradation were greater in columns amended with hydrogen peroxide
which provided the highest concentration of available oxygen (Brown et al., 1984). At concentrations greater than
100 ppm, hydrogen peroxide may degas to form air bubbles and block some of the pores in the aquifer.
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Decomposition of hydrogen peroxide may also be catalyzed by iron and fluctuations in pH (Britton and Texas
Research Institute, Inc., 1985). In addition to comprising part of the nutrient formulation, certain forms of
phosphate, such as potassium monophosphate, can be used to stabilize hydrogen peroxide solutions. To reduce
phosphate adsorption by the soil, a combination of simple and complex polyphosphate salts can be used (Brown et
al., 1986). Results from a field test in which hydrogen peroxide was used to increase the dissolved oxygen content
of the ground water indicated an increase from 1 to 15 ppm within 70 hr at a monitoring well located 25 ft
downgradient of the injection well.
Raymond et al. (1986) received a patent on a process for stimulating biodegradation of organic contaminants in
the subsurface with hydrogen peroxide. The patent described several formulations of nutrient and hydrogen
peroxide solutions and processes that can stabilize the decomposition of hydrogen peroxide, control movement of
the solution through the aquifer, remove metal ions from the subsurface which catalyze hydrogen peroxide
decomposition, and disrupt biofilms that form at the point of injection. Hydrogen peroxide decomposition can be
controlled by the addition of peroxidase, oxidase, or a transition metal (iron, copper, manganese, chromium, or
other material including the chelated forms of these metals). In addition, condensed phosphates can be perfused
into the aquifer to deactivate or remove substances that catalyze hydrogen peroxide decomposition.
Movement of the hydrogen peroxide solution through the formation can be controlled by hydratable polymeric
materials, interface modifiers, and densifiers (Raymond et al, 1986). The hydratable polymeric materials, such as
polysaccharides, polyacrylamides, and poly aery lam ide copolymers, increase the viscosity of the solution. An
increase in viscosity reduces the rate of diffusion and slows the movement of the solution. The addition of
surfactants will decrease the interfacial tension, prevent clays from swelling, disperse materials throughout the
zone of contamination, and decrease the metal-catalyzed decomposition of hydrogen peroxide. For example, the
zone of treatment can be extended into the capillary zone by adding soluble orthophosphoric salts and condensed
phosphoric acids to increase the capillary rise of the aqueous solution. Salts, such as sodium chloride, calcium
chloride, and sodium bromide, can be used to change the density of the nutrient solution. Biofouling can be
controlled by adding high concentrations (0.5 to 3%) of hydrogen peroxide; the effectiveness of hydrogen
peroxide in controlling biofouling may be enhanced by the addition of dilute acid.
Hydrogen peroxide has been used to enhance oxygen supply in the subsurface in many remedial programs. In
most cases, the additional oxygen was required to degrade hydrocarbons in aquifers contaminated from gasoline
spills. One case study involved the contamination of a relatively impermeable soil (ground water movement, 2 to
3 ft/year) with gasoline (Yaniga and Mulry, 1984). About 50 to 60 percent of the free product was recovered;
however, concentrations of hydrocarbon in the range of 3,700 to 7,200 ppm remained sorbed to the soil. A
feasibility study was conducted to identify an in situ microbial population capable of degrading the hydrocarbons
when supplied with nutrients and oxygen. Hydrogen peroxide was used as the source of oxygen. After two
months of operation, free product recovery reached a maximum of 25 to 30 gal/day, numbers of hydrocarbon-
degrading bacteria increased one to three orders of magnitude, and the concentration of sorbed product fell to a
range of 2300 to 2900 ppm.
In another in situ biorestoration program designed to clean up gasoline from a leaking underground storage tank,
oxygen was initially supplied by air stripping and sparging and then by hydrogen peroxide. A layer of heavy silt
loam which was underlaid by a layer of fractured shale and silt stone was contaminated by the spill (Yaniga,
1982). The gasoline infiltrated into the ground water and the resulting plume, containing dissolved hydrocarbons,
contaminated 12 domestic water wells with concentrations ranging from less than 10 ppb to 15 ppm. Ground
water was withdrawn and an air stripper was used to remove volatile organics and add oxygen; the oxygenated
water was then recirculated into an infiltration gallery to facilitate removal of the trapped organics. Air stripping
reduced the dissolved organics to less than 0.1 ppm. Additional oxygen was introduced by sparging air through
the 20 to 30 foot water column in the wells. The initial dissolved oxygen levels in the contaminated zone were 0-
1 mg/L, whereas those in the uncontaminated wells were 7-9 mg/L. Dissolved oxygen levels rapidly increased in
the periphery of the plume; after six weeks of air sparging, the concentration of dissolved oxygen increased to 3-5
mg/L in the contaminated zone. The nutrient solution included ammonium chloride, sodium phosphate, and
various mineral salts. Existing monitoring wells were used to add the nutrients because nutrient diffusion was
slower than desired. In the first 20 months of treatment, reductions of 50 to 85 percent of the dissolved
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hydrocarbons were achieved; however, treatment was continued because significant levels of hydrocarbon
remained. After this treatment period, reductions in the concentrations of dissolved hydrocarbons were minimal
and believed to be a result of inadequate oxygen supply (Brown et al., 1986); in addition, a biofilm had developed
and plugged the injection wells (Yaniga et al., 1985b).
To increase the concentration of dissolved oxygen, a trial experiment using hydrogen peroxide was conducted in
which 5 gallons of 100 ppm hydrogen peroxide was added to an injection well 40 ft from a pumping well (Yaniga
et al., 1985b). As a result, the dissolved oxygen content in the ground water collected from the pumping well
increased from 0.5 to 8 ppm in 24 hr with a concommitant increase in microbial activity. The decision was then
made to add 100 ppm hydrogen peroxide to the infiltration gallery and to the injection wells to increase the
concentration of dissolved oxygen in the formation. Addition of hydrogen peroxide also controlled growth of the
biofilm on the screens of the injection wells. After addition of hydrogen peroxide, the concentration of
hydrocarbons in the ground water had been reduced from 15 to 2.5 ppm. Continuation of this treatment removed
the dissolved hydrocarbons in ground water from eight of the 12 wells. Between 200 and 1,200 ppb remained in
the other four wells (Brown et al., 1986).
Hydrogen peroxide-assisted biodegradation followed by granular activated carbon polishing was used to treat a
spill of waste solvents and fuel (Brenoel and Brown, 1985; Westray et al., 1985). The source of the contamination
was an excavated area around several leaking tanks at a laboratory facility. The fill material and soil surrounding
the storage tanks was contaminated with 1000 to 3000 ppm hydrocarbon. The hydrocarbon was composed of
xylenes, benzene, toluene, ethyl benzene, C4-C12 alkanes, and C7-C10 alkanes. About 2,600 L of free product was
recovered using a sump pump; however, an estimated 1,100 to 3,400 L of the hydrocarbon remained. The
subsurface consisted of sand and sandy clay with ground water flows in the hundreds of feet per year. The total
number of bacteria in the well water ranged from 300,000 to 420,000 cells/ml; hydrocarbon degraders ranged
from 5,400 to 6,100 cells/ml. Carbon adsorption and enhanced bioreclamation were considered for remedial
action. Carbon adsorption was estimated to require 10 to 20 years and cost $470,000 to $850,000 whereas
enhanced bioreclamation was estimated to require four to eight months and cost $180,000 to $270,000 (Brown et
al., 1986). Enhanced bioreclamation was chosen and the process design consisted of four injection wells and a
pumping well with flows of 15 to 25 gpm (Westray et al., 1985). A nutrient solution consisting of ammonium
chloride and sodium phosphates was injected by batch addition. Hydrogen peroxide was injected continuously
following a short period during which only nutrients were added; addition of nutrients without oxygen had little
effect because of the initial low dissolved oxygen content (0.8 ppm). After 72 days, 1,200 pounds of nutrients and
250 gallons of the hydrogen peroxide solution had been added. The number of hydrocarbon degrading
microorganisms increased 130 fold and the concentration of dissolved oxygen in the ground water increased to
10.5 ppm after the biorestoration program was initiated. The concentration of hydrocarbon decreased from 22,700
to 581 ppb in 44 days and to non-detectable levels in one monitoring well after 72 days. Elevated concentrations
of contaminants detected in another monitoring well were thought to result from a leaking line. An estimated 150
to 400 gallons of the mixed fuels and solvents had been degraded (Brown et al., 1986). However, the formation
became partially clogged after 72 days of operation. Clogging of the formation may have resulted from the
movement of silt and degradation of the cement that lined the storage tank vault (Brenoel and Brown, 1985). An
activated carbon system was then used as a polishing step to reduce the hydrocarbon concentration below 10 ppb
in the tank vault and soil.
The cost for a 6 to 18 month bioreclamation program at the laboratory facility was estimated between $180,000
and $270,000 (Brown et al., 1986). Estimates were $50,000 to $75,000 to start the bioreclamation process and
$130,000 to $220,000 for services and nutrients. The cost for excavation was estimated between $600,000 and
$1,500,000, and the program was projected to take less than six months; however, facilities on the site would
restrict excavation. Withdrawal and treatment by carbon absorption was estimated to cost $470,000 to $850,000
and require 10 to 20 years because of limited extractability of the contaminants.
A less successful demonstration of enhanced bioreclamation using hydrogen peroxide was reported by Brown and
Norris (1986). A formation consisting of silt, sand, and gravel deposits was contaminated by a spill of 80,000 gal
of unleaded gasoline. Two subsurface zones were identified in the test area: 1) a fine quartz sand with some
limestone and dolomite grains and ferro-magnesium minerals with traces of limonite and pebbles of dolomite,
39
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limestone, and granite and 2) another zone of fine quartz sand with large amounts of fines and silt which impeded
ground water flow. The hydraulic conductivity ranged between 8.8 to 15.2 x 104 cm/sec. A free product recovery
program was implemented; however, between 300 and 10,000 ppm of hydrocarbon remained in the soil and 50 to
60 ppm remained in the ground water after five years. The concentration of total hydrocarbons in the cores
averaged 5,477 ppm with a range of 4,823 to 6,331 ppm for several groups. The highest concentrations were
detected at the water table at depths of 24 to 26 ft. The treatment zone was estimated to contain 16,500 % 6,700
(std. dev.) pounds of gasoline. Ambient nutrient levels in the ground water were less than 1 ppm and the
dissolved oxygen content was less than 0.4 ppm. Total counts and counts of hydrocarbon-degrading bacteria
grown on nutrient agar were 1.2 x 103 and 2 x 102 cells/g, respectively. Biostimulation was tested in a section of
the plume surrounded by two triangular patterns of monitoring wells which surrounded an inner infiltration
gallery. Nutrients were batch fed to the inner gallery and then followed by addition of hydrogen peroxide
solutions which were gradually increased from 0 to 500 ppm. Phosphorus levels reached 100 to 250 ppm in the
inner gallery and ranged from 1 to 10 ppm outside the gallery. The concentration of nitrogen ranged from 100 to
250 ppm in the inner gallery and from 10 to 50 ppm outside of the gallery. Total bacterial counts reached 106
cells/ml in the inner gallery and 6.0 x 103 cells/ml outside of the gallery; gasoline-utilizers also increased.
The concentration of hydrocarbon in the soil was measured after 0,32,91 and 164 days during the test and at
depths of 23,25,27, and 29 ft below the land surface (Brown and Norris, 1986). During the test, the
concentration of hydrocarbon was reduced from 5,490 to 1,874 ppm (65%). Removal of hydrocarbons was
highest (from 5,643 to 1,743 ppm) in the inner gallery near the injection area whereas a low permeability zone
was less effectively treated because of reduced circulation of nutrients. In addition, the concentration of total
hydrocarbons was reduced by 63 percent outside and between the galleries. The hydrocarbon concentration at the
water table was reduced from 6,087 to 4,058 ppm; from 2,946 to 1,008 ppm immediately below the water table.
The data indicate that substantial quantities of hydrocarbons remained adsorbed onto the soil after in situ
biostimulation although more improvement may have occurred with continued treatment.
A field demonstration of in situ, biorestoration using hydrogen peroxide in a very gravelly clay loam was adversely
affected by the low permeability (3.9 x 10'5 to 3.3 x 10'3 cm/sec) of the soil (Lee et al., 1986). The heterogeneity
of the soil and distribution of the contaminants made it difficult to inject nutrients and pump water. The
contamination resulted from a disposal pit containing chromium sludges, electroplating wastes, chlorinated
solvents, cresols, chlorobenzenes, and other compounds (Wetzel et al., 1985a; Wetzel et al., 1985b). The organic
compounds that were identified included tetrachloroethylene, trichloroethylene, trans- 1,2-dichloroethylene, and o-
and p-dichlorobenzene. Heavy metals present at concentrations greater than 10 mg/L included antimony,
chromium, copper, lead, nickel, and zinc and the concentrations of silver, cadium, and mercury were high in some
locations (Wetzel et al., 1985a and Wetzel et al., 1985b). The formation consisted of gravel lenses and layers of
fine grained soils with low hydraulic conductivities. The water table was perched, only 4 to 8 feet thick, and
exhibited seasonal fluctuations. Direct microbial counts in soil ranged from 7.6 to 170 x 106 cells/g (wet weight);
viable counts ranged from less than 100 to 7 x 106 cells/g on both rich and poor media. Laboratory studies
conducted under aerobic conditions indicated that the chlorobenzenes, hydrocarbons, and aromatics could be
biodegraded. Degradation of the organic contaminants separated by gas chromotography and thought to represent
n-alkanes was faster in the aerated microcosms than in those supplied with hydrogen peroxide, which may
indicate hydrogen peroxide toxicity. Unresolved hydrocarbons representing branched alkanes were removed
under aerobic but not anaerobic conditions. The results from these microcosm studies suggested that biological
degradation was feasible, but the heterogeneity of the subsurface and the contaminants present seriously limited
application of the biorestoration process. The presence of heavy metals was not expected to prevent
biodegradation, but the treatment process could induce metal mobilization.
The treatment system design for in situ biorestoration consisted of nine extraction and four injection wells that
were connected to a central surge tank and a distribution box (Heyse et al., 1985). One upgradient and two
downgradient wells were installed to monitor the influence of the treatment in untargeted areas. Nutrients were
added two weeks before the hydrogen peroxide. After two months of treatment, the effectiveness of the treatment
could not be determined because of a change in analytical methods; however, a number of problems were noted
with the field demonstration. Hydrocarbon levels increased in the ground water for unknown reasons. In
40
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addition, the nutrient solution precipitated initially when added on a continuous basis; the precipitation problem
was reduced by switching to batch amendments. The temperature in the well water increased and fluctuated
because the impermeable soil would not sustain flow adequate to prevent the pumps from overheating. Microbial
numbers in the infiltration zone remained low, perhaps due to hydrogen peroxide toxicity. Antimony and lead
may have been mobilized within the aquifer and the nutrients had not reached most of the wells at the time of this
report (Science Application International Corporation, 1985).
Continued treatment for six months resulted in decreases in the levels of chlorobenzene and total hydrocarbons
(Wetzel et al., 1986). The breakthrough of the nutrient solution was rapid in the highly permeable zones, but poor
in the less permeable strata. Elevated concentrations of carbon dixoide were detected in the treatment zone which
suggested an increase in microbial activity. However, the concentration of most of the organic compounds did not
decline and biorestoration of the site was not successful.
Ozone--In addition to hydrogen peroxide, ozone (03) can be used as an alternate source of oxygen. Ozone was
used in an in situ biorestoration program to remediate a hydrocarbon spill in a railroad yard in Karsruhe, West
Germany (Nagel et al., 1982). The presence of organic contaminants in the drinking water wells for the city of
Karlszruhe was traced to the hydrocarbon spill in the train yard. The concentration of organics, iron, and
manganese in the ground water increased but the dissolved oxygen content decreased. The water was withdrawn,
treated with 1 g of ozone/gram of dissolved organic carbon for four minutes and then reinjected into the formation
through five infiltration wells at a rate of 80-120 m3/hr. Supplemental nutrients were not added. The purified
water formed a barrier to prevent further contamination of the withdrawal well. The ozone treatment increased
the dissolved oxygen levels in the ground water which stimulated the microbial population and enhanced the
degradation of the contaminants in the aquifer. The maximum efficiency of introducing dissolved oxygen into the
ground water was 80 percent of the initial concentration of ozone. Oxygen consumption by the indigenous
microbes reached approximately 40 kg/day. The dissolved organic carbon decreased from a range of 2.5 to 5.5 g/
m3 to a steady state value of slightly more than 1 g/m3; little mineral oil hydrocarbons remained. The levels of
iron and manganese were also reduced. Although the total number of bacterial cells increased, microbial counts
on media which selected for disease-causing organisms did not increase. The removal of the hydrocarbons
probably resulted from both in situ microbial activity and chemical oxidation by the ozone. Hydrocarbons could
not be detected in the biostimulated section of the aquifer in water collected 1.5 years after treatment.
Soil venting-Soil venting or air flooding can be used to supply oxygen for in situ biorestoration. Organic vapors
from the unsaturated zone are removed by increasing the flow of soil gases using vapor recovery wells and air
inlet wells (Crow and Minugh, 1986). The volatile organic contaminants partition into the soil gas and are
transported to the vapor recovery wells. The increase in soil gas flow in the unsaturated zone makes more oxygen
available to reaerate the ground water. Wilson and Ward (1986) suggested that air flooding can be used to supply
oxygen during in situ biorestoration. Air contains 20 times more oxygen than water and is less viscous. For a fine
sand or silt, about 32,000 pore volumes of water in comparison to 4000 volumes of air is required to meet the
oxygen demand for degradation of saturating concentrations of hydrocarbons. Fewer volumes of each is required
for more porous soils. In addition to supplying oxygen, air flooding also removes vapors by physical weathering.
In the absence of a layer of pure product floating on the water table, the water table can be lowered by
withdrawing ground water to bring the contaminated region into the unsaturated zone for treatment. However,
lowering the water table would produce large quantities of contaminated ground water that must be treated. Soil
venting is currently being applied for in situ restoration of gasoline-contaminated soil by supplying oxygen to a 95
foot thick unsaturated zone where the contaminants are held (Kuhlmeier and Sunderland, 1986).
Collodial gas aphrons-Michelsen et al. (1985) suggested that a colloidal dispersion of air contained in a surfactant
matrix could be used to supply oxygen for in situ bioreclamation. The microdispersion of air, known as colloidal
gas aphrons, is prepared by passing air or pure oxygen through a venturi with a very small gas entry port into a
surfactant solution or by use of a spinning disk apparatus. The resulting colloidal material is basically a
suspension of fine soap bubbles with diameters of 25 to 50 microns that contain up to 65 percent gas. Up to about
55 percent of the pore space in sands can be filled with the colloidal air dispersion. Coarse sand was better than a
fine sand in retaining the colloidal air. Laboratory tests indicate that the technique can support the aerobic
metabolism of phenol and hexadecane; better removal of hexadecane was achieved when the suspension was
41
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prepared with oxygen (90%) than air (70%). Methods for application of the colloidal air in the field are still in the
devlopmental stages. In addition, biodegradation of the surfactants has also not been addressed.
Summary of Aerobic lit Situ Biostimulation Processes
There are a number of advantages and disadvantages in using in situ biorestoration (Table 2-4). Compounds
ranging from petroleum hydrocarbons to solvents have been treated by in situ biorestoration (Table 2-5). Unlike
many aquifer remediation techniques, in situ bioreclamation can often treat contaminants that are sorbed to soil or
trapped in pore spaces. In addition to treatment of the saturated zone, organics held in the unsaturated and
capillary zone can be treated when an infiltration gallery or soil flushing is used. Biodegradation in the subsurface
can be enhanced by increasing the concentration of dissolved oxygen, through the use of hydrogen peroxide,
ozone, or a colloidal dispersion of air (colloidal gas asphrons). Complete biodegradation (mineralization) of
organic compounds usually produces carbon dioxide, water, and an increase in cell mass. However, incomplete
degradation (biotransformation) of organic materials can produce byproducts that are more toxic than the parent
molecule. An example of biotransformation is the degradation of isopropanol to acetone at a hazardous waste site
described by Flathman and Githens (1985). The levels of acetone increased initially, but declined after most of
the isopropanol was removed. In situ biorestoration may rely on the biodegradation potential of the indigenous
Table 2-4. Advantages and Disadvantages of Biorestoration (J. R. B. Associates, 1982; Yang and Bye, 1979)
Advantages
• Can be used to treat hydrocarbons and certain organic compounds, especially water-soluble pollutants and
low levels of other compounds that would be difficult to remove by other methods
• Environmentally sound because it does not usually generate waste products and typically results in complete
degradation of die contaminants
• Utilizes the indigenous microbial flora and does not introduce potentially harmful organisms
• Fast, safe and generally economical
• Treatment moves with the ground water
• Good for short-term treatment of organic contaminated ground water
Disadvanta ges
• Can be inhibited by heavy metals and some organics
• Bacteria can plug the soil and reduce circulation
• Introduction of nutrients could adversely affect nearby surface waters
• Residues may cause taste and odor problems
• Labor and maintenance requirements may be high, especially for long term treatment
• Long term effects are unknown
• May not work for aquifers with low permeabilities that do not permit adequate circulation of nutrients
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Table 2-5. Contaminants Treated By In Situ Biostimulation
Contaminants
Treatment Description
References
high octane
gasoline
gasoline
gasoline
gasoline
gasoline
gasoline
gasoline
gasoline
unleaded gasoline
mineral oil
hydrocarbons
gasoline
air sparging with nitrogen
and phosphorus addition
air sparging with complete
mix of inorganics
air sparging with addition of
complete inorganic nutrient
solution
air sparging and addition of
nutrients
dissolved oxygen supplied by
an air stripper and sparging;
nutrients also added
dissolved oxygen supplied by
an air stripper
hydrogen peroxide plus
nutrients
initial treatment utilized air
stripping; hydrogen peroxide
used later with the nutrient
formulation
hydrogen peroxide supplied
the oxygen
withdrawn water treated with
ozone and reinfiltrated
soil venting used to supply
oxygen to unsaturated zone
Raymond etal., 1975
Raymond et al., 1975
Jamison et al., 1975
Jamison et al., 1976
Raymond et al., 1978
Minugh etal., 1983
Suntech, 1978
Yanigaetal., 1985a
Yanigaetal.,1985b
Brown etal., 1985b
Yaniga and Mulry,
1984
Yaniga, 1982
Brown et al., 1985b
Yaniga etal., 1985b
Brown and Norris,
1986
Nagel et al., 1982
Kuhlmeier and
Sunderland, 1986
(Continued)
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Table 2-5. (Continued)
Contaminants
Treatment Description
References
waste solvents
and alkanes
nutrients plus hydrogen
peroxide
Brown etal., 1985b
Westray el al., 1985
Brenoel and Brown,
1985
Brown et al., 1986
methyl chloride,
n-butanol,
dimethyl aniline,
acetone
methylene
chloride
ethylene glycol
isopropanol and
tetrahydrofuran
aliphatic
hydrocarbon
plasticizer
chloroform
withdrawal and treatment by
an activated sludge process
and recharge of aerated
nutrient-laden water
withdrawal and treatment with
air stripping followed later
by treatment in an activated
sludge unit and recharge
treatment following withdrawal
with ethylene-degrading
bacteria and nutrients and
then recharge
treatment in an above ground
reactor with addition of
acclimated microbes to the
aquifer along with nutrients
activated sludge and recharge
of acclimated bacteria and
nutrients
activated sludge bioreactor
with the bacteria innoculated
into the subsurface
Jhaveri and
Mazzacca, 1983
Jhaveri and and acetone
Mazzacca, 1985
Quince, et al., 1985
Flathman et al., 1985
Flathman etal., 1985
Flathman and Caplan,
1985
Flathman and Githens,
1985
Winegardner and
Quince, 1984
Winegardner and
Quince, 1984
subsurface microflora which usually contains few pathogenic organisms unless the aquifer has been contaminated
with wastewaters (Keswick, 1984). The time required to treat subsurface pollution using in situ biorestoration can
often be faster than some withdrawal and treatment procedures. A gasoline spill in Ambler, Pennsylvania, was
remediated in 18 months using in situ biorestoration whereas pump and treat techniques were estimated to require
100 years to reduce the concentrations of gasoline to potable levels (Raymond et al., 1976). In situ biorestoration
can also cost less than other remedial options. Flathman and Githens (1985) estimated that the cost of in situ
biorestoration would be one-fifth of that for excavation and disposal of soil contaminated with isopropanol and
44
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tetrahydrofuran and in addition would provide an ultimate disposal solution. The areal zone of treatment using
biorestoration can be larger than other remedial technologies because the treatment moves with the plume and can
reach areas which are otherwise inaccessible.
There are also disadvantages to in situ biorestoration programs. Many organic compounds in the subsurface are
resistant to degradation. In situ biorestoration requires an acclimated population; however, adapted populations
may not develop for recent spills or recalcitrant compounds. Heavy metals and toxic concentrations of organics
may inhibit microbial activity and preclude the use of the indigenous microflora for in situ biorestoration at some
sites. One option in this instance would be to remove the inhibitory substances and then seed the subsurface with
appropriately adapted microorganisms; however, the benefits to adding microorganisms to the subsurface are still
undemonstrated. The formation and injection wells may clog from profuse microbial growth which results from
the addition of oxygen and nutrients. In one biostimulation project, microbial growth produced foaming in the
well casings (Raymond et al., 1976a). In addition, the hydrodynamics of the restoration program must be properly
managed. The nutrients added must be contained within the treatment zone because the profusion of inorganics
into untargeted areas can result in eutrophication. High concentrations of nitrate can render ground water
unpotable. Metabolites of partial degradation of organic compounds may impart objectionable tastes and odors.
For example, the incomplete degradation of gasoline under low dissolved oxygen conditions resulted in phenol
production; phenol was then degraded when more aerobic conditions were achieved (Raymond et al., 1978).
Biostimulation projects require continuous monitoring and maintenance for successful treatment; whether these
requirements are greater than those for other remedial actions is debatable. The process results in increased
microbial biomass which could decompose and release undesirable metabolites. In addition, microbial growth can
exert an oxygen demand that may drive the system anaerobic and result in the production of hydrogen sulfide or
other objectionable byproducts. The long term effects of biorestoration are unknown. In situ biorestoration is
difficult to implement in low permeability aquifers in which perfusion of nutrients and oxygen is slow or
negligible; however, many in situ physical and chemical remediation processes are subject to the same
restrictions. The success of m silu treatment schemes in low permeability aquifers depends on transporting the
nutrients to the microflora or the active agent to the contaminants. The process has been used in different
hydrogeological formations (Table 2-6).
Potential problems for any aquifer restoration program include reversible adsorption of the contaminants, poor
delineation of the plume, inadequate sizing of the recovery system, pollution at depth, high costs, treating and
disposing of large amounts of pollutants, constraints on ground water pumping, access to the contaminated area,
and substantial quantities of pollutants in the vadose zone (Schmidt, 1983). To decrease the expense of an aquifer
cleanup, Nyer (1985) advocated a policy of life cycle design for remedial actions in which some of the equipment
could be recycled and used at other sites. An example of this system was proposed to remediate contaminated
ground water from a Gulf Coast hazardous waste site. The ground water contained high concentrations of phenol
and enough dissolved solids (15,000 mg/L) to be considered a brine. The treatment system consisted of two
activated sludge units, a fixed film-activated sludge unit, a dual media filter, and a carbon adsorption column.
The components of the treatment system could be easily changed to accomodate the change in concentration of
the contaminants during the clean up process.
Innovative Processes
There are a number of innovative, generally unproven, processes that potentially can be applied to in situ
biorestoration. These processes include land treatment, techniques that decrease the surface tension to enhance
the mobility and improve the biodegradability of the contaminants, application of enzymes, and treatment beds.
Land treatment-Land treatment is a process in which the indigenous microflora in surface soils degrade the
organic material contained in the soil. Loehr and Malina (1986) suggested that land treatment is useful for
disposal of organic wastes from municipal sludge, petroleum, wood preserving, leather tanning, coal gasification/
liquefaction, food processing, and pulp and paper production. Land treatment involves the addition of the organic
waste to the soil, mixing to aerate and incorporate the organics into the soil and, if needed, adding fertilizer to
stimulate microbial activity. The process must be carefully managed to prevent overloading the assimilative
capacity of the soil and to prevent migration of the inorganic nutrients, organics, and heavy metals (Ross and
Phung, 1983). Major factors that control biodegradation in land treatment are listed in Table 2-7. Land treatment
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Table 2-6. Types of Aquifers Where In Situ Biostimulation Has Been Utilized
Aquifer Description
Flow Characteristics
Reference
high permeability
dolomite
medium to coarse sand
alluvial fan deposit of
sand, gravel, and cobbles
with some clay and silt
poorly sorted mixture
of boulders, pebbles,
cobbles, sand, silt and
clay
perched water table in
unstratified, unsorted
layer of clay, silts,
sands, gravels, and
cobbles above a clay
layer
tank vault filled with
pea gravel surrounded
by sand and sandy clay
strata
glacial outwash composed
of silt, sand, and
gravel
coarse sands and gravel
shale and siltstone
coarse sand with greater
than 5% gravel
pumping rate of 265
to 378 L/min
pumping rate of 65
to 151 L/min
flow of 2.4 m/day
hydraulic conductivity
of9.4xlO-5to
1.7xlO-3cm/sec
pumping rate of 38 to
to 57 L/min
flow rate in excess
100 m/yr
pumping rate of 151
L/min
hydraulic conductivity
of 8.8x10-" to
1.5xlO-3cm/sec
hydraulic conductivity
of 2.1 cm/sec
pumping rate of 68 L/min
gradient of 0.015 to
0.02 m/m; flow of 0.61
to 0.91 m/yr
(Continued)
Raymond et al., 1976
Raymond etal., 1975
Jamison et al., 1975
Jamison et al., 1976
Raymond et al., 1986
Minugh etal., 1983
Jhaveri and Mazzacca,
1983
Jhaveri and Mazzacca,
1985
Quince et al., 1985
Flathman et al., 1985
Westray et al., 1985
Brown etal., 1985b
Brenoel and Brown,
1985
Brown and Norris, 1986
Nagel et al., 1982
Brown etal., 1985b
Yaniga et al., 1982
Yanigaetal.,1985b
Yaniga and Mulry, 1984
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Table 2-6. (Continued)
Aquifer Description Flow Characteristics Reference
glacial till composed of Yaniga et al., 1985b
sand, gravel, and boulders
in a silty clay matrix
connected to a fractured
sandstone
shallow basin containing flow of 27 to 38 L/min Flathman and Githens,
sand and pea gravel 1985
Table 2-7. Factors That Control Biodegradation in Land Treatment (Huddleston et al., 1986)
1. Chemical structure of the waste
2. Presence of appropriate numbers of microorganisms capable of degrading
the wastes
3. Concentration of the wastes
4. Supply of oxygen
5. Optimal water content of between 25 and 85% of the water holding capacity
6. Optimal temperatures between 20 and 30°C
7. Optimal pH levels between 6 and 8
8. Availability of inorganic nutrients, principally nitrogen and phosphorus
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may be advantageous in comparison to other remedial techniques because it requires minimal operation and
maintenance and is a proven technology for some wastes (Reible and Wetzel, 1983). However, the process may
result in incomplete destruction of the organic wastes, the soil may be difficult to aerate effectively, the wastes
must be contained within the treatment zone, air pollution may result, and large areas of land are required. Land
treatment is usually limited to the upper 1.5 m of soil (Ryan et al., 1986) which restricts its use in aquifer
remediation. However, land treatment may be used to treat excavated soil or wastes that are concentrated at the
water table of shallow aquifers. Law Engineering Testing Company (1982) recommended that land treatment of
ground water contaminated with gasoline be considered when a suitable site is available. In comparison to other
treatment technologies, land treatment was rated highly on the basis of effectiveness, capital costs, reliability, and
operability (Kebe and Brockman, 1984).
Land treatment was used to treat a spill of 1.9 million liters of kerosene (Dibble and Bartha, 1979). About 200 m3
of soil was excavated and the contaminated ground water was withdrawn and treated. The contaminated soil was
treated by adding lime and fertilizer (nitrogen, phosphorus, and potassium in the ratio of 10:1:0.85) and frequently
tilling the soil to a depth of 46 cm. The fastest rate of kerosene degradation occurred during the warmer months
of July and August. The concentration of kerosene was reduced from 0.87 percent to innocuous levels in the
upper 30 cm of soil during a 21 month period. However, kerosene persisted at a depth of 30 to 45 cm, perhaps a
result of reduced aeration. Within the first seven months, most of the n-alkanes and unresolved hydrocarbons
were degraded. A test for phytotoxicity after land treatment indicated that the phytotoxicity of the kerosene had
been reduced but not completely eliminated; crop yields were 20 percent below those in a control area 23
months after the spill.
Techniques that reduce the interfacial tension-Insoluble organic compounds that are sorbed to soils can be
mobilized and made more available for microbial attack by decreasing the interfacial tension between the
compounds and water. The interfacial tension can be decreased with dispersants, surfactants, extractants, and
emulsifiers. Dispersants have been used in remediation programs to control marine oil spills with some success
(Colwell and Walker, 1977). Addition of disperants can increase the rate of reaction, but may not increase the
extent of hydrocarbon degradation. However, not all dispersants enhance degradation of hydrocarbons and some
may be toxic to microorganisms. Mulkins-Phillips and Stewart (1974a) reported that only one of four dispersants
stimulated biodegradation of crude oil by marine bacteria; however, all four dispersants caused shifts in the
microbial population.
The addition of surfactants to mobilize organics sorbed to soils has been tested in laboratory studies. A
combination of nonionic and ionic surfactants was most effective in removing gasoline from sand columns by
simple displacement and by draining the capillary zone of the gasoline (Texas Research Institute, 1979). Some of
the surfactants identified in this study were biodegradable whereas others exhibited varying degrees of toxicity.
Ellis et al. (1984) demonstrated that surfactants could remove up to 95 percent of the crude oil and
polychlorinated biphenyls trapped in sand columns whereas aqueous washes failed to remove appreciable
quantities of these contaminants. Surfactants may be used in combination with biorestoration to remediate aquifer
contamination problems. A surfactant wash can mobilize the residual hydrocarbon in the unsaturated zone and
render trapped hydrocarbon in the saturated zones more available for biodegradation (Wilson and Ward, 1986). A
surfactant which is biodegradable and non-toxic is required. The application of surfactants to subsurface
contaminants may present additional environmental problems by spreading contaminants to sections of the aquifer
previously uncontaminated.
Emulsifers can be used to increase the surface area and render the oil more degradable (Atlas, 1977). Emulsifiers
can be either chemical additives or biological agents. Robichaux and Myrick (1972) reported that one chemical
emulsifier increased the microbial decomposition of oil eighteen-fold; however, other emulsifiers were less
successful and many may have been toxic. Broderick and Cooney (1981) reported that emulsifiers are produced
by a variety of organisms in freshwater environments, especially those associated with sediments. Laboratory
studies conducted by Vanlooke et al. (1978), showed that 10-20 percent of the oil adsorbed to soil was removed
after the addition of a nutrient solution containing ammonium nitrate and peptone; microbial metabolities were
thought to be responsible for the enhanced desorption. The ground water microflora in an aquifer contaminated
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with aviation fuel was reported to emulsify hydrocarbon when supplied with dissolved oxygen, nitrogen and
phosphorus (Ehrlich et al., 1985). Micelles and microemulsions of the hydrocarbon are likely to be formed by
bioemulsifiers which may facilitate transport of the hydrocarbon into the cell (Wilson, 1986). Biosurfactant-
producing bacteria may be used to remediate contaminated aquifers but their use will be controlled by: 1) the
physical and chemical characteristics of the contaminant, 2) the geophysical and geochemcial characteristics of
the formation such as pore size distribution and permeability, water quality, and oxygen concentration, and 3)
competition with indigenous microflora. The hydrocarbon that is mobilized as a result of bioemulsification may
be withdrawn from the aquifer and treated by above ground techniques (Wilson et al., 1986). The microbial
conversion of the hydrocarbons to more polar compounds such as alcohols, ketones, phenols, or organic acids will
also mobilize the contaminants.
Another application of bioemulsifiers is enhanced oil recovery. Clark et al. (1981) found several aerobic
microbial species that could be used to bioemulsify oil in situ. Zajic and Akit (1983) found two bacterial strains
that produced high concentrations of surfactants; one could remove bitumen from tar sands when grown on
hexadecane. A bacterial culture supported on molasses was able to release 19.5 to 48.7 percent and utilize 2 to 51
percent of the oil in a formation within 10 days (Lazar, 1983). Field trials were successful in two of seven
reservoirs, increasing yields by 10 to 200 percent.
Extraction techniques such as steam flooding, alcohol flooding, and thermal flooding also may be used to
mobilize organic contaminants in the subsurface; however, they have not been demonstrated in the field (Wilson
and Conrad, 1984). Horizontal or vertical water sweeps can be used in permeable aquifers to reduce the quantity
of hydrocarbons before treatment by other methods such as ia situ biorestoration.
Enzymes as an innovative treatment technique-Another innovative in situ process is the addition of enzymes to
degrade specific organic compounds. In one investigation, a parathion hydrolase enzyme isolated from a mixed
culture ofPseudomonas was added to wet and dry plots of soil amended with the organic phosphorus insecticide
diazinon (Paulson et al., 1984). In both wet and dry plots, removal was initially faster in the enzyme-amended
soil than in the control; however, diazinon levels in the test and control plots were similar after 408 hours. The
effectiveness of an enzyme depends upon its stability in the enviromnment and contact with the substrate.
Adequate mixing to insure contact may be difficult to achieve in an aquifer. In addition, enzymes may be better
substrates for microbial metabolism than many organic pollutants. The stability of an enzyme in the environment
may be adversely affected by changes in pH and solute concentrations.
Treatment Beds-Treatment beds are another innovative process currently under development. The process
consists of a trench which intercepts contaminated ground water and either a biological or chemical treatment bed
which removes the contaminants. Chemical treatment beds for organic compounds include activated carbon or
synthetic resins (Ehrenfeld and Bass, 1984). Biological treatment can be accomplished using processes similar to
trickling filters in which microorganisms colonizing a surface are supplied with oxygen and nutrients, if
necessary, and degrade the contaminants which enter the treatment bed. Permeable treatment beds may plug or
exhibit channeling, which reduces their effectiveness. Similar results could be obtained without the treatment bed
by implementing in situ biorestoration in a narrow zone that intercepts and contains the plume.
Potential for Anaerobic Processes
Anaerobic degradation pathways in the subsurface—Anaerobic processes are important in the subsurface
environment because oxygen may be depleted in contaminated aquifers as a result of aerobic microbial activity.
However,low levels of oxygen will support some microbial activity. Once the dissolved oxygen content in ground
water declines as a result of microbial activity, replacement depends on recharge, reaeration from soil gases, and
mixture with oxygenated waters surrounding the organic plume (Borden and Bedient, 1986; Borden et al., 1986).
Degradation of a variety of compounds under anaerobic conditions has been demonstrated to occur in aquifers and
laboratory experiments using subsurface materials. However, anaerobiosis may retard the degradation of many
compounds (Hutchins et al., 1985). The sequence of microbial processes that occur as environmental conditions
49
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change from aerobic to anaerobic in the subsurface usually follows the pattern of aerobic respiration,
denitrification, manganese and iron reduction, sulfate reduction, and finally methane formation (Bouwer, 1985;
Downes, 1985). Net energy production decreases as the redox potential decreases (Downes, 1985). Bouwer and
McCarty (1983a;1983b) demonstrated differences in the degradation of organic compounds under different redox
potentials; chloroform and 1,1,1-trichloroethylene were degraded by methanogenic, but not denitrifying bacteria.
Ehrlich et al. (1982; 1983) reported the degradation of phenolics, but not polynuclear aromatics such as
naphthalene, under methanogenic conditions. Recently Kuhn et al. (1985) documented removal of
tetrachloroethylene, the xylene isomers, and dichlorobenzene isomers under denitrifying conditions. Wilson and
Rees (1985) snowed that degradation of benzene, ethylbenzene, toluene, and o-xylene occurred in methanogenic
aquifer material from a landfill, although the process was slow compared to aerobic pathways. The concentration
of toluene had been reduced by 87 percent after six weeks, however, more than 20 percent of the benzene,
ethylbenzene, and o-xylene added to the microcosms persisted beyond 40 weeks. In the same study,
trichloroethylene and styrene degraded under anaerobic conditions, whereas chlorobenzene persisted. Suflita and
Gibson (1985) reported that 13 of 19 halogenated isomers of benzoate, phenol, and phenoxyacetate persisted at
concentrations greater than 90 percent of that initially added to subsurface materials collected from a sulfate-
reducing zone; however, only 3,4-dichlorobenzene remained at concentrations greater than 5 percent of that
originally added to methanogenic samples collected downgradient of the sulfate reducing zone. Maximal numbers
of sulfate-reducing and methanogenic bacteria are found at redox potentials of -100 to -150 and -250 to -350 mV,
respectively (Van Engers, 1978). Halogenated aliphatics such as trichloroethylene, tetrachloroethylene, carbon
tetrachloride, and 1,1,1-trichloroethane can be mineralized or dehalogenated under reducing conditions (Parsons et
al., 1985) to potentially more toxic compounds such as vinyl chloride (Vogel and McCarty, 1985; Wood et al.,
1985).
Anaerobic processes in in situ biostimulation-Anaerobic processes may be of potential use in in, situ
biorestoration processes. The redox potential would be selectively adjusted to favor the degradation of a
particular contaminant. In addition to adjusting the redox potential, the pH of the ground water could be adjusted
to the neutral or alkaline conditions required for sulfate reduction, methanogenesis, and usually denitrification.
Anaerobic degradation of organic compounds would probably require less inorganic nutrient supplementation
because less energy and therefore biomass is produced (Rittman and Kobayshi, 1982). Batterman (1983) added
nitrate to ground water contaminated with hydrocarbons in an attempt to promote denitrification. The
contaminated aquifer consisted of an 8 to 10 meter thick layer of sand which contained some silt and clay beds
and a ground water flow of 4 m/day. The water was withdrawn from a deeper uncontaminated aquifer, aerated,
passed through a sand filter, and amended with nitrate at 300 mg/L before being recharged to the shallow aquifer.
Phosphate was not added because it was not limiting. The authors suggested that anaerobic degradation accounted
for the removal of 7.5 tons of hydrocarbon within a period of 120 days. Removal of 1 mg of the hydrocarbon
required 3.3 mg of nitrate (Batterman and Werner, 1984). The concentration of aliphatics declined slowly from
1.5 to about 0.7 mg/L whereas the total aromatics declined from 5.5 mg/L down to about 1.5 mg/L in
approximately one year. The rate of decline in the concentration of xylene was much slower than that of benzene
and toluene. Water was injected during the test which resulted in a rise in the level of the hydrocarbons as well as
the water table into the unsaturated zone. There was an overall 40 percent reduction in the concentration of
hydrocarbon as a result of the treatment process. Insufficient information was provided to determine if anaerobic
degradation was responsible for the removal of the contaminants or if the removal was due to the oxygen
introduced when the injection water was aerated before it was recharged into the shallow aquifer.
Degradation of low concentrations of organic compounds under methanogenic conditions, with acetate added at
higher concentrations as a primary substrate, has been demonstrated (Bouwer, 1985). McCarty (1985) proposed a
scheme to treat contaminated ground water anaerobically using the primary substrate concept. The system
consists of an above ground reactor to which substrate and nutrients are added, a well casing bioreactor which
operates anaerobically like a trickling filter, and the aquifer. The above ground reactor is used to develop an
acclimated population. The effluent from the above ground reactor is injected into the well casing bioreactor to
introduce acclimated microbes into the aquifer or enhance adaptation of the indigenous population to the
contaminants. Once the acclimated population has devloped, use of the above ground reactor can be discontinued.
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A method that utilizes sequential aerobic and anaerobic conditions to degrade hazardous wastes has been studied
in soils and may be applicable to subsurface cleanup. An insecticide, methyoxychlor, was slightly degraded in
soil under either aerobic or anaerobic conditions after three months of incubation. When the samples were
converted from an anaerobic to an aerobic status, mineralization of the methoxychlor increased 10 to 70 times of
that observed in soils maintained aerobically throughout the incubation period (Fogel et al., 1982). The
enhancement in methoxychlor degradation in soils exposed to anaerobic and then aerobic conditions may be a
result of dechlorination of the insecticide under anaerobic conditions and degradation of the dechlorinated
products under aerobic conditions. This anaerobic-aerobic treatment scheme may be useful in biorestoration of
aquifers contaminated with halogenated compounds. The aquifer could be managed like a sequencing batch
reactor in which an acclimated population is exposed to deoxygenated water, then to aerobic conditions and then
the treated water is withdrawn. The hydraulically managed system is then allowed to sit idle until the next cycle
is initiated.
Rates of degradation under anaerobic conditions are typically slower than those under aerobic conditions; in
addition, organic compounds may not be mineralized under anaerobic conditions even after long periods of
incubation (Wilson and Rees, 1986). However, anaerobic treatment may be required to degrade pollutants that are
recalcitrant under aerobic conditions; also, anaerobic treatment may require less management. The application of
anaerobic conditions to biorestoration is still in the development stage and more research is required to
demonstrate its usefulness in the field.
3. Addition of Specialized Microbial Populations to the Subsurface
In addition to stimulating the indigenous microbial population to degrade organic compounds, another innovative
but not yet demonstrated technique is to add microorganisms with specific metabolic capabilities (Lee and Ward,
1985b). Specialized organisms may be inoculated into the subsurface environment or the environment may be
altered to favor growth of a population with specific metabolic capacities. Populations that are specialized in
degrading target compounds are selected by enrichment culturing or genetic manipulation. Enrichment culturing
involves exposure of microorganisms to increasing concentrations of a contaminant or mixture of contaminants.
The type of microorganisms that is selected or in essence, acclimates to the contaminant, depends on the source of
the inoculum, the conditions used for the enrichment, and the substrate (Atlas, 1977). Acclimation can result from
an increase in the number of organisms that can degrade the contaminant, new metabolic capabilities that result
from genetic changes, or an increase in the quantity of the enzymes necessary for the transformation (Spain et al.,
1980). The genetic changes include overproduction of enzymes, inactivation or alteration of regulatory gene
control, or production of enzymes with altered specificities (Ghosal et al., 1985).
Genetic manipulation of microorganisms to produce specialized populations that can degrade target contaminants
is a relatively recent development. According to Kilbane (1986), genetic engineering may accelerate and focus
the process of evolution. Genetic manipulation can be accomplished by two different methods. In the first
method, the organisms are exposed to a mutagen such as ultraviolet light, nitrous oxide, or 8-azaquinonone and
then a population with specialized degradative capabilities is isolated by enrichment culturing (Zitrides, 1978;
Kopecky, 1982); however, this may produce weakened strains because the process is non-specific and affects the
entire genome (Zitrides, 1978). In the second method, recombinant DNA technology is used to change the genetic
structure of the microorganism (Kilbane, 1986). The genetic structure is changed by inserting a DNA fragment,
often a plasmid that codes for a specific degradative pathway, into another organism. A plasmid is a piece of
DNA that exists independently from the cell's chromosomes (Birge, 1981). The extra-chromosomal DNA can be
transformed from one bacterium to another by conjugation, transduction, or transformation. Multiple degradative
capabilities can be placed on a single plasmid that will allow the organism to degrade an array of compounds or
complete the degradation of a nonbiodegradable molecule. Genetic engineering can be used to stabilize the
degradative traits coded by the plasmid, increase the number of plasmids in a cell, amplify enzyme production and
activity, invoke multiple degradative traits, or produce a novel degradative pathway (Pierce, 1982). In addition,
organisms with different substrate affinities, pH optima, or degradation rates can be fashioned (Johnston and
Robinson, 1982a).
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Genetic Engineering to Enhance Degradative Activity--
Genetic engineering has been used to enhance the degradation of the recalcitrant pesticide, 2,4,5-
trichlorophenoxyacetic acid (2,4,5-T). Biodegradatiori of the pesticide is usually very slow (Kilbane et aL, 1982).
A mixed culture of microorganisms that uses 2,4,5-T as a sole carbon and energy source was obtained by a
technique called plasmid-assisted molecular breeding (Kellog et al., 1981). The technique involves inoculating a
chemostat with microorganisms from a variety of hazardous waste sites and organisms that carry an array of
plasmids that code for degradation of specific xenobiotics. A pure culture that could use 2,4,5-T as a sole carbon
and energy source was isolated from the mixed population and tentatively identified as Pseudomonas cepacia
(Kilbane et al., 1983). In addition, the culture, designated P. cepacia AC1100, was reported to oxidize many
chlorophenols. Degradation of both 2,4-dichlorophenoxyacetic acid (2,4-D) and 2,4,5-T was expressed in another
strain of £. cepacia after conjugal transfer of two plasmid from an Alcaligenes entrophs sp. that degraded some
chlorinated phenoxy herbicides (Ghosal et al., 1985). An inoculum of 2 x 107 cells/g of £. £§gacja AC1100
degraded 95 percent of the 1,000 mg/L 2,4,5-T added to soil at 25 percent moisture and incubated at 30°C
(Chatterjee et al., 1982). Less 2,4,5-T was removed with a smaller inoculum size and different temperatures and
moisture contents. In addition, the 2,4,5-T degrading bacteria did not survive in soil without 2,4,5-T or when the
concentration of the compound had been depleted (Kilbane et al., 1983). Field trials to determine the
effectiveness of the 2,4,5-T-degrading bacteria have not been conducted.
Colaruotolo et al. (1985a) received a patent for "microbial degradation of obnoxious organic wastes into
innocuous materials." The process involves isolation of microbial cultures from samples of soil and leachate
from a hazardous waste site by enrichment culturing and then application of the purified strains in the field to
remove the contaminants. Microorganisms capable of degrading selected isomers of chlorotoluene,
dichlorotoluene, and dichlorobenzoate were isolated. Conjugation and transformation experiments were
conducted to transfer the plasmid DNA, which conferred the ability to degrade some chloroaromatics, from the
original isolates to another organism. The patent claimed that the organisms could be used to decontaminate soil,
remove contaminants in the air, mineralize toxic organics in the leachate from a chemical landfill and thereby
reduce the concentrations of noxious chemicals.
Issues in Genetic Engineering of Microbes-
Organisms that can not easily exchange their genetic information with other organisms and are restricted to
growth under defined environmental conditions are preferred candidates for genetic manipulation (Pierce, 1982).
Issues concerning the use of genetically engineered organisms in the environment include: 1) adverse effects on
human health, 2) how to effectively monitor their dispersal, 3) survival of the engineered organism in the
environment, 4) regulation of activity in nontarget areas; and 5) determination of set risk levels acceptable to the
public (Joyce, 1983). Many scientists argue that the engineered organism is not radically different from that
which is genetically unaltered. The release of genetically engineered organisms into the environment is of great
concern and some time may elapse before these organisms are used (Fox, 1985). The survivability of genetically
altered organisms in the environment is also of concern. Surrogates of genetically engineered organisms which
carried antibiotic resistance were added to samples of sewage, lake water, and soil and survived at rates that
varied with the strain and environment tested (Liang et al., 1982). Some of the antibotic-resistant strains reached
steady-state concentrations in lake water and sewage; however, all strains declined in the soil after a period of one
month. Pseudomonas strains that degrade 2,4-dichlorophenol and p-nitrophenol were isolated from soil by
enrichment culturing techniques. The ability of the isolates to degrade the phenol derivations was variable when
inoculated into lake water, sewage, and soil (Goldstein et al., 1985).
Inoculation of a specialized microbial population into the environment may not produce the desired results for
many reasons (Table 2-8). The concentration of the target compound required to support activity of a specific
degrader may be limiting. Toxic or antimicrobial substances such as antibiotics may be found in many
environments. High density inocula may be grazed by predators and the degradative capacity severely decreased
if the growth rate of the introduced organisms is slow. In addition, adequate mixing to ensure contact of the
organism with the pollutant will be difficult to achieve in the subsurface.
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Table 2-8. Reasons Why Introduced Organisms Fail to Function in the Environment (Goldstein et al., 1985)
1. The concentration of the compound is too low
2. The environment contains some substance or organisms that inhibit growth or activity, including predators
3. The inoculated organism uses some other organic other than the one it was selected to metabolize
4. The organic is not accessible to the organism
Most hazardous waste sites involve contamination of the environment with more than one compound. Therefore a
mixture of organisms may be necessary to degrade all of the compounds in the waste (Atlas, 1977). Populations
that have adapted to degrade many organic contaminants may be isolated from biological treatment processes
such as sewage treatment, which receive pollutants. The efficacy of an inoculated population of specific
degraders will depend on environmental constraints such as temperature, pH, and the concentrations of substrate,
nutrients, and oxygen (Adas, 1977; Zajic and Daugulis, 1975). Successful results from inoculation of foreign
organisms are more likely in simple environments because the environment can be controlled more easily. An
example of inoculation into a simple environment would be the introduction of bacteria into a biological reactor,
oil tanker ballast tanks, or fermentator; these also provide the benefit of containing the microorganisms. To avoid
problems encountered with inoculation of foreign organisms into the environment, samples from the contaminated
environment can be collected, microorganisms that can degrade the pollutants can be cultured by enrichment
techniques or genetically engineered, and finally the specialized population can be reintroduced into the
environment from which they came (Omenn, 1986). In addition, genetic manipulation of oligotrophic bacteria
with high affinity enzyme systems may be advantageous because these enzyme systems will allow the organism to
attack low concentrations of organic pollutants (Johnston and Robinson, 1982b).
Seeding Aqueous Environments with Microorganisms—
Inoculants of specialized microorganisms have been used in treatment of contaminated water. Atlas and Bartha
(1973) tested several commercial bacterial preparations and found that the inocula were ineffective in treating oil
spills in the marine environment. However, the addition of fertilizer and a bacterial seed isolated from an
estuarine environment increased petroleum degradation in a saline but not in a freshwater pond (Atlas and
Busdosh, 1975). After six weeks, 50 percent of the oil remained in the saline pond. The lack of activity in the
freshwater pond suggests that the inoculum should be cultured from an environment similar to that being treated.
Colwell and Walker (1977) suggested that seeding would be unsuccessful in environments such as the ocean;
however, contained spills and lagoons may be amenable to such treatment. Gutnick and Rosenberg (1977) stated
that "there is no evidence to support the claim that "seeding" oil slicks with microorganisms reduces oil
pollution by stimulating petroleum biodegradation."
Seeding Soil Environments with Microorganisms-
The efficacy of inoculating soil with acclimated bacteria to remove selected contaminants was tested in a series of
experiments (Wetzel et al., 1981) using experimental chambers set up in greenhouses. The contaminants, aniline
and formaldehyde, were added to three types of soils (clay, sandy loam, and organic-rich) and plants were seeded
in the chambers. Removal of the contaminants by a mixed microbial population from primary sewage effluent
and an acclimated population was investigated. Formaldehyde was not removed in organic soils amended with
sewage and acclimated bacteria; however, this treatment was successful in the upper and middle zones of the sand
and clay soils. Aniline was removed in the organic and sandy soils after a second application of sewage
microorganisms, nutrients, and yeast extract. Chemical oxidation of the organics using hydrogen peroxide was
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effective in reducing aniline concentrations. None of the treatments were successful in removing aniline from the
clay soil. The removal of chlordane and 2,4-dinitrophenol by mutant adapted microbial cultures was also
investigated. The inoculum was successful in degrading 2,4-dinitrophenol from the upper layer of the clay soil
only. The authors suggested that the sewage inoculum was a low cost, effective method for removal of aniline
and formaldehyde in most soil types; however, addition of the adapted population was not successful in these
tests.
Inoculation of soils to remove chlorinated organics and pesticides has been attempted. Daughton and Hsieh
(1977) reported that inoculation of sterilized soils with a parathion-acclimated culture reduced the concentration
of the insecticide by 85 percent; however, the efficiency of the inoculum in non-sterile soil was greatly reduced.
Focht and Brunner (1985) used an Acinetobacter strain as an inoculum to degrade biphenyl and polychlorinated
biphenyls in soils. The inoculum increased the initial and maximum mineralization rates and the disappearance of
the more heavily chlorinated biphenyls, but the overall extent of mineralization of polychlorinated biphenyls was
not greater than that in inoculated soil to which biphenyl had been added. The process was thought to be a
cometabolic-commensal metabolism of the PCBs.
Remediation of soil contaminated with hydrocarbons by inoculating with hydrocarbon degrading organisms has
been met with varying success. Schwendinger (1968) demonstrated that inoculation of a hydrocarbon- degrading
strain of Cellumonas in soil contaminated with petroleum increased the rate of reclamation in comparison to soils
amended with only nutrients. Jobson et al. (1974) reported that the application of 106 cells of oil-degrading
bacteria per cubic cm of soil slightly increased the degradation of the C^- to C^- group of n-alkanes in
comparison to soils amended with fertilizer only. However, Lehtomaki and Niemela (1975) reported that
brewer's yeast added to soils served primarily as a fertilizer rather than as an inoculum to actively degrade the oil.
Seeding boreal soil with an oil-degrading inoculum increased microbial activity (Hunt et al., 1973). In laboratory
studies, the addition of 300 ppm nitrogen and 100 ppm phosphorus, inoculation, and adjusting the pH to 7,
increased microbial activity by at least a factor of four in comparison to unamended samples after 40 days of
incubation. An increase in plant growth in an oil contaminated area in response to fertilizer addition was shown in
field studies; however, the increased growth could have resulted from the addition of fertilizer or enhanced
removal of the petroleum. In contrast, Westlake et al. (1978) reported no beneficial effects from the addition of
oil-degrading bacteria to boreal soils. The lack of enhancement may be a result of inadequate application of the
inoculum. The type of organisms isolated from enrichment culturing depends on conditions used during the
isolation procedure. For example, enrichments made at 4 and 20°C contained different organisms, and cultures
enriched on a low quality crude were better adapted to utilize a lower quality crude than cultures enriched on a
high quality crude (Jobson et al., 1972). These data suggest that enrichments for specialized populations should
be conducted using the environmental conditions and contaminants that are unique to the site under investigation.
An inoculum of pentachlorophenol-degrading organisms has been used to decontaminate soil, river water, ground
water and other freshwaters (Martinson et al., 1984). A Flavobacterium sp. that could mineralize
pentachlorophenol (PCP) was isolated from a man-made channel which was exposed to the compound for several
weeks (Crawford and Mohn, 1985). In addition to mineralizing PCP, the microorganism could attack a number of
other chlorinated phenols but not all isomers (Steiert and Crawford, 1985). The Flavobacterium sp. at a cell
density of 106 cells/ml removed over 90 percent of the PCP added to river water, ground water and other fresh
waters, usually within 48 hours (Martinson et al., 1984). The organisms ability to degrade PCP was best between
15 and 35°C, and at pH values between 7.5 and 9.0. Inoculum densities as low as 104 cells/ml resulted in efficient
removal of PCP. The time required to remove the PCP increased with increasing concentrations of PCP. When
added to uncontaminated soil, the PCP was rapidly mineralized (Crawford and Mohn, 1985). The highest extent
of mineralization occurred in soils with moisture contents between 15 to 20 pecent.
Mineralization of PCP was observed at inoculum densities as low as 3.1 x 103 cells/g; however, a slightly higher
extent of mineralization was observed at a cell density of 3.1 x 106 cells/g (Crawford and Mohn, 1985).
Mineralization of PCP in one uninoculated soil began after seven days of incubation and mineralization proceeded
to the same point as the sample inoculated with 107 cells/g. Concentrations of PCP in soil contaminated from a
wood treating landfill were reduced from 298 to 58 ppm after four applications of the inoculum in a period of 100
days. In another contaminated soil, PCP levels were reduced from 321 to 41 ppm after one application of seed,
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but similar levels of removal were observed in the uninoculated control. The seed could not remove PCP from a
third soil in which the concentration of PCP had been diluted 10 fold to 553 ppm and the pH adjusted to
neutrality. Addition of 10s cells/g soil of a culture of PCP-degrading Arthrobacter sp. reduced the half-life of PCP
from two weeks to 15 hours (Finn, 1982). Edgehill and Finn (1983) reported that the rate of PCP disappearance
was proportional to inoculum size that ranged from 10* to 106 cells/g soil. Up to 85 percent of the PCP was
removed within 12 days in soil in which the seed had been thoroughly mixed; however, only 50 percent was
removed in the unmixed soil. Brown et al. (1986) suggested that fixed film reactors with a PCP-adapted
population may be used to treat waters contaminated with PCP at concentrations below the threshold of toxicity.
A consortium that was attached to rocks from an artificial stream amended with PCP was generally able to
degrade PCP as fast as the Flavobacterium sp. described by Crawford and Mohn (1985). A treatment system
using two fixed film reactors in series was then proposed; the first reactor would reduce high concentrations of
PCP and the second reactor would contain organisms that could remove PCP to low levels. The consortium was
able to remove PCP to less than 1 mg/L when the initial concentrations were less than 1 mg/L.
Seeding the Subsurface with Microorganisms--
Inoculation of bacteria into the subsurface for biorestoration has been met with some success, but the contribution
of the introduced bacteria to the overall cleanup can not be readily determined. In most cases, the role of the
introduced bacteria in degradation of the contaminants can not be determined because appropriate control plots
were not incorporated into the experimental design and the results were not quantitatively measured throughout
the course of the project. The biggest concern of inoculation into the subsurface is ensuring contact between the
specialized cells and the target contaminants. The cells may be filtered out of the perfusing solution or sorbed
onto soil before reaching the contaminants (Bouwer, 1984). In addition, normal die-off may control the
movement and spread of bacteria in well-sorted sand, gravels, fractured rock, and karstic limestone.
Microbial movement through the subsurface depends on the characteristics of the soil and microorganisms. Only
1 percent of an inoculum of a Pseudomonas strain passed through a 2-inch sandstone core after washing with 123
pore volumes (Jenneman et al., 1984). Penetration of bacteria into sandstone cores with hydraulic conductivities
greater than 100 millidarcies was rapid; however, penetration in cores with hydraulic conductivities below 100
millidarcies was slow (Jennemen et al., 1985). Motile bacteria moved three to eight times faster than nonmotile
bacteria. Hagedorn (1984) summarized the results of selected studies on the maximum distance that
microorganisms moved in various soils: 19.8 m in 27 weeks in a fine sand; 10.7 m in a sand and sandy clay in
eight weeks; 24.4 m in a fine and coarse sand (time of travel not reported); 30.5 m in a sand and pea gravel
aquifer in 35 hours; 0.6 to 4 m in a fine sandy loam (time of travel not reported); 457.2 m in a coarse gravel
aquifer in 15 days; 28.7 m in 24 to 30 hours in a crystalline bedrock. Bacteria have moved as far as 920 meters in
the subsurface at rates up to 350 m/day (Gerba, 1984). Microbial movement through soil macropores is an
important mechanism of transport in all subsurface soils except sandy soils and those that are disturbed (Smith et
al., 1983).
Transport of microorganisms in the subsurface can occur. However, in situ biorestoration programs using
inoculation techniques will be affected by adverse conditions that decrease the survivability of microorganisms in
the environment. Several factors must be considered before an in situ biorestoration program utilizing acclimated
bacteria is implemented. The source, quantity, nature and biodegradability of the contaminants, and the
environmental conditions of the site must be determined (McDowell et al., 1980). In addition, laboratory tests to
determine the kinetics of degradation, the potential for inhibition under various conditions, requirements for
oxygen and nutrients, and the effects of temperature should be conducted. The formation must be permeable
enough to perfuse nutrients and the inoculum through the zone of contamination.
Aquifer Remediation Using Inoculation Techniques--
Inoculation of microorganisms into the subsurface has been used in aquifer remediation in conjunction with
wastewater treatment processes. These cases are summarized in Table 2-9. A representative system is shown in
Figure 2-3. In one case study, 7,000 gallons of acrylonitrile was spilled in a metropolitan area from a leaking rail
car (Walton and Dobbs, 1980). The receiving aquifer contained significant amounts of silt and clay and hence
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was rather impermeable. Initial treatment involved withdrawal and treatment of the ground water by air stripping.
After the concentration of acrylonitrile had declined to nontoxic levels, mutant bacteria were seeded into the soil.
The concentration of acrylonitrile declined from 1,000 ppm to nondetectable levels (limit of detection 200 ppb)
within one month; however, the role of the bacterial seed in acrylonitrile degradation could not be determined.
Table 2-9. Summary of Aquifer Remediation Case Histories Utilizing Introduced Organisms
Compound
Treatment Description
Reference
acrylonitrile
phenol and
chlorophenol
ethylene glycol
and propyl acetate
dichlorobenzene,
dichloromethane,
and trichloroethane
unidentified organic
compounds
formaldehyde
mutant bacteria added after
concentrations had been
reduced by air-stripping
initial treatment by
adsorption onto GAC
followed by innoculation
with mutant bacteria
treatment above ground
and later with specialized
bacteria
initial treatment with
air stripping and then
innoculation with a hydro-
carbon-degrading bacteria
hydrocarbon-degrading
bacteria added after levels
reduced by GAC and air
stripping
commercial degrader added to
above ground treatment system
formed from rail ballast
Walton and Dobbs, 1980
Walton and Dobbs, 1980
Qunice and Gardner
1982aandb
Quince and Gardner,
1982aandb
Ohneck and Gardner,
1982
Sikes et al., 1984
Quince and Gardner (1982a; 1982b) documented the cleanup of 100,000 gallons of various organic compounds,
including ethylene glycol and propyl acetate, over a 250,000 square foot area. The soil consisted of a thick silty
clay that extended to a depth of more than 50 feet; migration of the organics into the main aquifer was prevented
by the structure of the formation. Containment and recovery of the organics were limited to the perched water
table located in the upper clay layer. The contaminated ground water was withdrawn and treated by clarification,
aeration, and granular activated carbon. A biostimulation program with specialized bacteria, nutrients, and air
was initiated after the levels of the contaminants had decreased from 2,000-10,000 ppm to less than 200 ppm.
During treatment, the concentration of ethylene glycol was reduced from 1,200 to less than 50 mg/L, propyl
acetate was reduced from 500 mg/L to less than 50 mg/L, and the total concentration of spilled compounds
declined from 36,000 to less than 100 mg/L. The resulting concentrations of contaminants were acceptable to the
regulatory agencies.
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DIRECTION OF GROUND WATER FLOW
Outside Water Source
To Specified
Discharge
INJECTION SYSTEM
ZONE OF CONTAMINATION
it
Contaminated
Water
T T T T T T T T T T
1 1 1 1 I 1 1 I I
RECOVERY SYSTEM
Pneumatic Recovery
Pump
f
Figure 2-3. Combination of above ground treatment with in sjlu biorestoration.
Quince and Gardner (1982a; 1982b) documented the cleanup of a number of organic chemicals including
dichlorobenzene, methylene chloride, and trichloroethane that contaminated the subsurface as a result of a spill
from vacuum system, soil flushing, air stripping, and then inoculation of commercial hydrocarbon-degrading
bacteria into an above ground reactor followed by recharge of the effluent into the subsurface. A commercial
decreased the concentrations of the organic contaminants after 36 hours of exposure. The operation was
terminated after a 95 percent reduction in the organic levels was achieved. The injected hydrocarbon degraders
were expected to complete the biodegradation in situ: however, the role of the added bacteria was not
demonstrated.
An accidental spill of 130,000 gallons of organic chemicals entered a 15 foot thick shallow unconfined aquifer
and resulted in total contaminant levels as high as 10,000 ppm (Ohneck and Gardner, 1982). A drinking water
aquifer was separated from the contaminated zone by 50 to 60 feet of silty clay. The contaminated ground water
was withdrawn and treated by clarification, granular activated carbon adsorption, and air stripping. A program to
enhance in situ biological degradation was initiated after the concentration of the organics had declined from as
high as 10,000 ppm to 1,000 ppm. The results of laboratory tests indicated that the indigenous bacteria could
degrade the contaminants when supplied with nutrients. Application of a commercial bacterial inoculum did not
increase the biodegradation rates of the organics; in fact, one compound (unidentified) was degraded slower by the
commercial hydrocarbon-degrading inoculum than the indigenous population. Effluent from the treatment system
was amended with hydrocarbon degrading bacteria, air, and nutrients, and injected into the vadose zone. As a
result, the concentrations of the contaminants in one soil core were reduced from 800 to 150 mg/L in two months.
In another area, the concentration of the chemicals in composited soil samples declined from 24,000 to 2,000 mg/
L. The concentrations of the organics in the ground water were reduced to less than 1 ppm, which met regulatory
approval.
57
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Incorporation of biological treatment into the restoration program decreased the cost of operation and
maintenance. The role of the commercial inoculum in the removal of the contaminants could not be determined;
in addition, laboratory studies indicated that the inoculum did not enhance biodegradation.
A spill of 20,000 gallons of a 50 percent solution of formaldehyde from a railroad tank car contaminated the soil
and railroad bed in Ukiah, California (Sikes et al., 1984). Contaminated surface and ground waters were removed
by a vacuum truck and 250 cubic yards of soil were excavated. Approximately 13 million gallons of water was
collected. The water was initially treated with hydrogen peroxide to reduce the concentration of formaldehyde
from 30,000-50,000 to 500-1,000 ppm by oxidation. See Section II.D. for more details of the use of hydrogen
peroxide in this case study. The feasibility of in silsi biological degradation of the remaining formaldehyde using
a commercial bacterial inoculum was then investigated. A commercial inoculum that contained specially cultured
microorganisms was chosen for the project. The biological treatment system consisted of a portable aeration tank,
a spray system, and a trickling filter. The ground water was heated to increase the destruction rate and the pH was
adjusted as necessary with sulfuric acid or soda ash; nitrogen and phosphorus were added as needed. The
inoculum was rehydrated with chlorine-free water and added to the system at a rate of 3 Ibs per day. The
concentration of formaldehyde in the treatment tank fell from greater than 700 to about 10 mg/L after 24 days.
The oxygen uptake rate in the sump ranged from 12 to 82 mg/L hr1 and from 29-51 mg/L hr1 in the ballast gravel.
The treatment program was temporarily suspended for a day and the system was flushed. During this period, the
concentration of formaldehyde increased greatly; however, a rapid reduction in formaldehyde levels followed.
The authors suggest that the removal of the formaldehyde was a result of biological activity, however, they
concede that proving the role of microorganisms in formaldehyde degradation would be difficult. In addition, the
role of indigeneous and inoculated bacteria in formaldehyde degradation could not be separated.
4. Enrichment of Specific Populations
A strategy often used in industrial microbiology is to search through nature for an organism with specialized
metabolic capabilities and then culture that organism in a fermentor to protect it from competition. However, this
strategy would be difficult to apply in the subsurface environment because the specialized population must be
competitive in addition to performing the desired transformation. Enrichment culturing techniques are often used
to isolate organisms with specialized metabolic capabilities. The same concept can be used to identify conditions
that favor the colonization of that environment by organisms with special traits.
The microbial utilization of pollutants as carbon and energy sources has already been discussed. This section will
emphasize the metabolism of pollutants by microorganisms enriched on other primary substrates.
Oxygenated water-table aquifers are often polluted with chlorinated organic solvents such as trichloroethylene
(Wilson and McNabb, 1983; Wilson et al., 1981). The ubiquity of these compounds in oxygenated ground water
may result from their resistance to microbial attack under aerobic conditions in the subsurface. However, more
recent work has indicated that incubation of soils or aquifer materials with methane, propane, or natural gas will
enrich for microorganisms that co-oxidize trichloroethylene and a variety of other halogenated organic compounds
(Wilson and Wilson, 1985; Fogel et al., 1986; Strand and Shipper!, 1986; Hensen et al., 1985; Henry and Grbic-
Galic, 1986; Wilson and White, 1986; U. S. EPA, 1984; and Hensen et al., 1986). This technique may be
applicable to in situ restoration of aquifers which are contaminated with chlorinated organic solvents.
Using gaseous aliphatic hydrocarbons as the feedstock for a forced co-oxidation is advantageous because they are
non-toxic, relatively inexpensive, and widely available in the form of natural gas, liquified petroleum gas, and
propane. However, they do not support anaerobic metabolism. In addition, if the gases are supplied inadvertently
at concentrations that result in the microbial depletion of the available oxygen, undesirable by-products such as
foul-smelling organics, soluble iron, or hydrogen sulfide should not be produced. The disadvantage to enriching
for specialized populations using gaseous aliphatic hydrocarbons is the explosion hazard of the hydrocarbons
mixed in air at unsafe concentrations. One constraint on in situ restoration programs is that the reagents must be
dissolved in the perfusion water to reach the zone of contamination; hydrocarbons and the oxygen required for
their metabolism are not very soluble in water.
58
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Wilson and White (1986) developed a general relationship that may be used to predict the extent of removal of a
chlorinated organic as a function of the metabolism of a given amount of hydrocarbon feedstock. The relationship
is as follows:
C/Co= e* (2-2)
where C = the final concentration of the halogenated organic to
be co-oxidized,
Co = the initial concentration of the halogenated organic,
h = the amount of hydrocarbon feed-stock to be consumed,
k = a utilization constant.
At present, there are limited data available to calculate utilization constants and the generality of the relationship
has not been widely tested. The equation may prove to be a powerful tool in the engineering design of in situ
biorestoration programs. Some utilization constants for trichloroethylene exist (Wilson and White, 1986). As a
paper exercise, the extent of removal of trichloroethylene at a number of critical engineering limitations, including
the solubility of methane, propane, and oxygen (24,62, and 40 mg/L respectively) and the oxygen content of well-
oxygenated ground water (taken to be 8 mg/L) was estimated. Equation 2-2 was used to estimate the
concentration of trichloroethylene that could be brought down to 5 |ig/L, the Maximum Contaminant Level
proposed under the Safe Drinking Water Act (U. S. EPA, 1984). To preview the economics of in situ
biorestoration, costs of the primary hydrocarbon feedstock were estimated at 25 cents/kg and the cost of oxygen
supplied as hydrogen peroxide was estimated at 400 cents/kg oxygen supplied.
The predictions of trichloroethylene degradation by populations supported on gaseous aliphatics using Equation 2-
2 are illuminating (Table 2-10). At ambient oxygen concentrations, the reduction in the concentration of
trichloroethylene supported on either propane or methane is environmentally insignificant. The reductions
supported by saturating concentrations of oxygen will probably be useful, but not sufficient to treat most
contaminated water in one cycle; the water will have to be circulated and reinjected with oxygen a number of
times, or the oxygen will have to be supplied as hydrogen peroxide. Propane is about three times more
soluble in water than methane, and considerably greater removals of trichloroethylene are possible using propane
as the feedstock. Finally, the cost of trichloroethylene biorestoration in situ can probably be attributed to the cost
of supplying the oxygen.
The quantities of methane and propane are calculated from the equation of Wilson and White (1986), assuming
utlization constants for trichloroethylene of 0.075 L water treated/mg methane consumed and 0.10 L water
treated/mg propane consumed, and assuming utilization constants for cis- and trans-1,2-dichloroethylene of 0.3 L
water treated/mg methane consumed.
Vinyl chloride and cj§- and £ans- 1,2-dichloroethylene commonly occur in ground water contaminated with
trichloroethylene, and probably result from the reductive dechlorination of trichloroethylene (Barrio-Lage et al.,
1986). Utilization constants for these compounds are not available. Wilson and White (1986) estimated from the
data of Fogel et al. (1986) that the constants are greater than 0.03 L water treated/mg methane consumed. These
higher utilization constants make aquifers contaminated with these compounds much better candidates for in situ
biorestoration (Table 2-10). Hydrogen peroxide will probably not be required to achieve adequate treatment. If
the contaminated water is pumped to the surface for treatment, the limited solubility of oxygen becomes much
less of a problem. The water can be exposed to any desired volume of air in a fixed-film bioreactor (Wilson and
White, 1986).
Table 2-11 summarizes the prospects for biorestoration of aquifers contaminated with specific halogenated
compounds. These data were compiled by comparing the relative rates of degradation of these compounds in a
variety of experimental systems to the rates of transformation of trichloroethylene, and cjs- and trans-1T2-
59
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dichloroethylene, then assessing the rates in light of the relationships portrayed in Table 2-10. Prospects are rated
"good" if hydrogen peroxide will not be required, "fair" if hydrogen peroxide is required, and "poor" if
environmentally insignificant removals can not be attained with or without hydrogen peroxide.
Table 2-10. Estimated Quantities of Oxygen and Methane or Propane Requered to Bring the Concentration of
Trichloroethylene, sis- or trans-1.2-Dichloroethylene or Vinyl chloride to 5 |j.g/L
Initial Concentration Methane Required Propane Required
TCE DCEorVC mg/L cents per mg/L cents per
Hg/L iig/L 1000 gal. 1000 gal.
2,500 62 5.9
1,000 53 5.0
250 39 3.7
100 29 2.7
30 24 2.5
17 11 1.0
11 10 0.9
7 2.2 0.2
5.8 2.0 0.2
5.0
6,700 24 2.5
1,000 18 1.7
100 10 0.9
9 2 0.2
5.0 0.0
Oxygen Required
mg/L cents per
1000 gal.
230
190
140
105
96
40
40
8
8
0.0
96
71
40
8
0.0
340
290
210
159
146
146
107
The relationship of Wilson and White (1986) does not presuppose an upper limit on the concentration of the
chlorinated contaminant; however, an upper limit obviously exists and toxicity effects have frustrated research in
this area. Workers at both Stanford University and R. S. Kerr Environmental Research Laboratory have isolated
mixed microbial populations from nature that could degrade trichloroethylene, only to lose the ability to degrade
the compound when the primary alkane oxidizer was isolated in pure culture. It is tempting to conclude that the
trichloroethylene degrader is not an alkane oxidizer. However, other possibilities exist. In certain mixed cultures
or microcosms of aquifers, trichloroethylene started to inhibit oxidation of the hydrocarbon feedstock at a
concentration of about 1,000 jo.g/L (S. Fogel personal communication, unpublished data of John Wilson). This is
60
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far below concentrations that produce toxicity in ordinary heterotrophs. Perhaps the effective toxicant was
trichloroethylene epoxide, rather than trichloroethylene itself, and the organism that oxidized the primary
hydrocarbon feedstock was protected in mixed populations by other organisms. This toxicity threshold must be
more carefully defined to aid in identifying contaminated ground water amenable to biorestoration. This work
should be done with mixed-cultures or microcosms, using systems that simulate the conditions in the subsurface
environment.
No technique to remediate environmental contamination is universally applicable. However, there should be
many contamination incidents where biorestoration through a forced co-oxidation is the technology of choice,
either alone or in conjunction with physical containment. Successful application of the approach will require an
adequate understanding of the philosophy of the biotransformation, and quantitative information on the nutritional
ecology of the active organisms.
Table 2-11. Prospects for Treatment of the Common Halogenated Organic Contaminants in Aquifers Through
Co-Oxidation Supported on Gaseous Alkanes
Compound
Pump
and Treat
Treat in the
Aquifer
References
Tetrachloroethylene
(PCE)"
Trichloroethylene (TCE)
Good
None
Fau-
cis- 1,2-Dichloroethylene
trans-1,2-Dichloroethylene
Vinyl chloride
direct utilization may be
possible
1,1 -Dicholoroethy lene
Carbon tetrachloride"
Good
Good
Good
Fair
None
Fogeletal., 1986
Hensenetal.,1985
Hensenetal.,1986
Wilson and Wilson,
1985
Fogeletal., 1986
Hensenetal.,1985
Hensenetal.,1986
Wilson and White, 1986
Fogeletal., 1986
Hensenetal.,1985
Hensenetal.,1986
Fogeletal., 1986
Hensen et al., 1985
Hensenetal.,1986
Fogeletal., 1986
Hartman et al., 1985
Fogeletal., 1986
Hensenetal.,1985
Hensenetal.,1986
(continued)
61
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Table 2-11. (Continued)
Compound
Pump
and Treat
Treat in the
Aquifer
References
Chloroform
Methylene chloride
direct utilization
may be possible
1,1,1 -Trichloroethane
(TCA)
1,1,2-Trichloroethane
(TCA)
Poor
to Good
Poor
Fair
Poor
Poor
Strand and Shippert, 1986
Hensenetal., 1985
Hensen et al., 1986
Henson et al., 1986
Flathman et al., 1985
Hensen et al., 1985
Hensen et al., 1986
Wilson and White, 1986
Hensenetal., 1985
Hensen et al., 1986
1,1-Dichloroethane
(TCA)
1,2-Dichloroethane
direct utilization may
be possible
1,2-Dibromoethane (EDB)
direct utilization
may be possible
Poor
Poor
Fair
Hensenetal., 1985
Hensenetal., 1986
Hensenetal., 1985
Hensen et al., 1986
Wilson and McNabb, 1983
Janssen et al., 1985
Hensen et al., 1985
Hensenetal., 1986
•Removal of carbon tetrachloride and tetrachloroethylene seen in the soil exposed to natural gas is probably an
anaerobic process, and not a direct result of alkane oxidation (personal communication Michael Henson, R. S.
Kerr Lab, U.S. EPA, Ada, Oklahoma).
F. Hydrologic Considerations and Mathematical Modeling of Biorestoration
1. Hydrologic Considerations
A number of methods have been reported in the literature for the containment of contaminated ground water
through hydraulic control or through injection-pumping networks of wells. Biorestoration of a contaminant plume
may involve the addition of nutrients such as dissolved oxygen or hydrogen peroxide or the addition of microbes
capable of degrading a particular waste. In order for such additions to be successful, it may be necessary to use
hydraulic controls to minimize the migration of the plume during the in situ treatment process. Thus, hydrologic
considerations cannot be neglected in the biorestoration of aquifers.
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Hydraulic control methods depend to a large extent on variability of aquifer hydraulic conductivities, background
velocities, and sustainable pumping rates. Typical patterns of wells which are used to provide hydraulic controls
include: 1) the injection-production pair, 2) a line of downgradient pumping or recharging wells, 3) a pattern of
injection-production wells around the boundary of a plume, and 4) the ' 'double-cell'' hydraulic containment
presented by Wilson (1984). Well systems can also be used to capture and withdraw entire zones of contaminated
water for treatment above ground.
Analytical equations and graphical solutions are available for estimating flow rates and limits of hydrodynamic
isolation under various boundary conditions. Numerical computer models of ground water flow and contaminant
transport are required when site geology is complex, heterogeneous, and anisotropic. A simple hydrodynamic
isolation system within a uniform flow field involves the placement of a recharge well of the same strength
upgradient from a pumping well. Standard equations describe the head h(x,y) as a function of pumping rate,
ambient flow rate, and transmissivity of the aquifer. The region of recirculation which connects the stagnation
points can be evaluated, and provides a measure of the capture zone of contaminated ground water (Bear, 1979).
Wilson (1984) presents a "double-cell" hydraulic containment system which utilizes an inner cell and an outer
recirculation cell, with four wells along a line bisecting the plume in the direction of flow. The method is more
efficient in terms of flushing times and recirculation rates than the single cell. The double-cell method provides
added flexibility and a back up system if pumps should fail in either system.
Ozbilgin and Powers (1984) described hydrodynamic isolation systems for several EPA hazardous waste sites.
Pumping wells and an upgradient recharge trench were successful in retarding the advance of a contaminated
plume at the site in Nashua, New Hampshire. They concluded that hydrodynamic isolation systems are generally
less costly and time-consuming than physical containment structures, such as slurry walls. Well systems are more
flexible in that pump rates and well locations can be altered as the system is operated over a period of time.
Shafer (1984) indicated that pumping-injection systems can be used 1) to create stagnation (no flow) zones at
precise locations in a flow field, 2) to create gradient barriers to pollution migration, 3) to control the trajectory of
a contaminant plume, and 4) to intercept the trajectory of a contaminant plume. However, the determination of
pumping rates to achieve a pollution control objective can be difficult. Thus, investigators have explored the
application of optimization theory to determine optimal pumping rates for creation of hydraulic controls.
Gorelick (1982) and Atwood and Gorelick (1985) focus on using linear programming (LP) methods to determine
the best containment strategy in combination with a ground water flow simulator. From a specified set of
potential well sites, the model approach selects the best wells and optimal pumping/recharge rate schedules to
contain the contaminant plume. Shafer (1984) advocates the use of non-linear programming combined with a
ground water flow model and an advective transport model. The optimization method is applied to examples for
determing stagnation points in a flow field and for steering the trajectory of a contaminant plume. Optimization
methods offer more efficient solutions than the typical trial and error approaches for exploring cleanup strategies.
However, nonlinear programming requires the flow and transport models to run during each iteration in which
new pumping rates are selected and tested in the overall performance index or objective function. In the present
case, optimization methods are complex and time consuming and may not offer any improvements over
simulation for the complex case of biorestoration alternatives.
Successful biorestoration alternatives at a particular site depend on the hydrologic and geologic characteristics of
the aquifer. If the contaminant plume is moving rapidly through a sandy-gravelly zone, then hydraulic controls
may be required to halt the advance of the plume and to provide injection points for added nutrients or oxygen.
Pumping out ground water and surface nutrient additions prior to reinjection may provide a more controlled input
to the biorestoration process.
For the case of slowly moving organic plumes in a silty sand aquifer, it may be hydrologically difficult to pump or
inject recharge waters at rates greater than 5 or 10 gpm. In such cases, large numbers of wells may be needed to
provide better "hydraulic spreading" of treated recharge water. Simple lines of wells upgradient or downgradient
63
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of the plume may not provide the required circulation, and wells within the plume are usually needed. Five spot
patterns (one injection surrounded by four pumping wells) provide a useful network for many cases.
In summary, hydraulic controls for the containment of ground water should be carefully considered for any site
where biorestoration is a viable treatment alternative. In particular, injection-pumping well networks offer
advantages for the creation of stagnation (no-flow) zones or for the control of the trajectory of a contaminant
plume. Once the plume has been controlled hydraulically, then application of additional nutrients, oxygen, or
microbes can be better controlled and evaluated in terms of biodegradation
efficiency.
2. Modeling Biorestoration
Mathematical modeling of biorestoration processes is useful in simulating cleanup progress and can provide
insights into the kinetics of the restoration process. Modeling of the hydraulics of die site may also aid in
designing the optimum injection and production system. Development of mathematical models of the
biorestoration process requires: 1) a description of the kinetics of biodegradation/transformation in the
subsurface; 2) a description of the abiotic processes controlling the transport and availability of the contaminant
and other required nutrients; and 3) an appropriate procedure for combining the processes and predicting the effect
of the biorestoration technique. Most attempts at quantifying the transport and removal of contaminants in ground
water have relied on a solution of the classical form of the advection-dispersion equation. The general form of
this equation is
= V • (DVC - vQ + (2-2)
dt
where
C = contaminant concentration
t = time
v = velocity vector
D = dispersion tensor
R. = chemical and biological reaction terms
A = the del operator
Solutions to this equation have been obtained using a variety of analytical and numerical methods. Thorough
reviews of these methods may be found in Anderson (1979), Bear (1979) and Javandel e't al. (1984). In this
section, mathematical descriptions of biodegradation kinetics are reviewed along with commonly used
descriptions of abiotic transport processes. Commonly used techniques for solving these equations are then
reviewed as well as some of the advantages and disadvantages of mathematical modeling.
3. Kinetics of Biodegradation
In situ biorestoration usually involves the addition of electron acceptors and nutrients to enhance the growth of
microorganisms present in the subsurface and consequently increase the rate of contaminant biodegradation. In
order to model the degradation process, relationships are needed which describe the kinetics of microbial growth
and consumption of added nutrients and electron acceptors. These relationships are then combined with Equation
2-2 to describe the movement and consumption of the contaminant and added nutrients. One of the most popular
relationships for describing the growth and decay of microorganisms and consumption of organic substrate was
originally proposed by Monod (1942) and modified by Herbert (1956). This model takes the following form:
64
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and
dS », S
(2-4)
dt r K + S
where
X = microbial concentration (mg/L)
\i = maximum specific utilization rate (I/day)
Y = microbial yield coefficient (g/g)
S = limiting substrate (mg/L)
b = microbial decay rate (I/day)
K = substrate half saturation constant (mg/L)
Growth is assumed to be a hyperbolic function of some limiting nutrient. Microbial decay is assumed to be a
constant independent of other environmental conditions. When several compounds are used simultaneously,
Equation 2-3 can be modified as
y f* V j_ f* Y j. f* If 4. C*
Ol A.-I + l^« Aj + x*j «j "*" vj JVR T \s^
where
Cj = limiting nutrient i
K. = substrate half saturation constant for nutrient i
When C.»K., the function (C/K. + C.) goes to 1.0 and has no effect on the growth of the microorganisms, but
when C.«K., the growth rate will be directly porportional to the concentration of nutrient i. These equations also
predict that as the concentration of nutrient i decreases, the net growth rate will approach zero and eventually
become negative. For a population to survive, the long term growth rate must be greater than or equal to zero;
consequently, the concentration of any limiting nutrient i may not fall below some minimum (C^) where
(2-6)
\iY-b
65
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This equation suggests that microorganisms may not be capable of degrading organic contaminants below this
concentration. McCarty (1985) has suggested the addition of a second non-harmful substrate to support the
growth of the organisms and allow degradation to below C^; a more readily degradable substrate could suppress
the degradation of contaminant of interest. Yoon et al. (1977) have presented a mathematical model for
simulating growth of a mixed microbial population on multiple substrates. For two substrates:
where
f
• - *
(2-7)
where
Si2 =
K, , =
substrates 1 and 2
half saturation constants for substrates 1 and 2
maximum specific growth rates for substrates 1 and 2
inhibition constants
This model can be further extended to describe growth on multiple substrates.
Some contaminants will not be used as a carbon and energy source by the microorganisms but are transformed.
Schmidt et al. (1985) have shown that transformation of these compounds is proportional to microbial population
and contaminant concentration (C) where
- uXC
(2-8)
Schmidt et al. (1985) used a logistic curve to describe the change in microbial concentration in a batch system and
developed a series of equations for describing change in contaminant concentration with time for differing initial
microbial and contaminant concentrations. Use of the logistic curve greatly simplifies the mathematical
computations but does not allow simulation of the effects of changing aquifer parameters such as the addition of a
second substrate. In aquifer restoration, simulation of the microbial population using Equation 2-3 and change in
contaminant concentration by Equation 2-8 may provide a more useful prediction.
Various workers have suggested that the kinetics of microbial growth, decay and consumption of organic
contaminants in the subsurface are best described by models which include terms for transport into attached
biofilms or microcolonies. It is well known that most microorganisms in the subsurface are attached to soil
particles (Harvey et al. 1984). This is thought to be due to the competitive advantage attachment gives a
microorganism at low substrate concentrations (Heukelekian and Heller, 1940).
Workers at Stanford University have developed a series of models for simulating degradation of organics in
biofilms. The basic model assumes that degradation within the biofilm can be described by Monod kinetics.
Mass transport into the biofilm is by diffusion alone. The diffusive flux (J) is calculated from Pick's second law
J= -
(2-9)
66
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where
Sf = concentration of rate limiting substrate
z = coordinate orthogonal to biofilm
D = diffusivity
Different diffusivities are assigned to the biofilm and an effective diffusion layer adjacent to the biofilm.
Williamson and McCarty (1976) originally developed this model to simulate either substrate or oxygen limited
biodegradation in wastewater treatment biofilms. Rittman and McCarty (1980) modified this approach to describe
the steady state biofilm surrounding an injection well receiving tertiary treated wastewater. Bouwer and McCarty
(1984) have further expanded the approach to allow simulation of secondary utilization of trace organics while the
biofilm is supported by an undifferentiated chemical oxygen demand (COD). Most recently, Kissel et al. (1984)
have employed the biofilm concept to model carbonaceous oxidation, nitrification and denitrification within a
mixed culture biofilm.
Much of the work on biofilms at Stanford occurred as an outgrowth of field studies on organic degradation near an
injection well. In this region, substrate fluxes and ground water velocities will be high and a biofilm can be
expected to develop. Actual biofilms are rare in most aquifers and the majority of the microorganisms are present
as microcolonies. Molz et al. (1986) have recently modified the biofilm concept to describe the growth and decay
of microorganisms present in microcolonies. An average colony radius and thickness is used to describe the
microcolonies. Growth and decay within the colony is simulated by Monod kinetics and includes both oxygen
and substrate limitation. Transport to the microcolony is limited by a diffusion layer at the colony surface.
4. Modeling Subsurface Transport
Keely et al. (1986) present a concise overview of evolving concepts of subsurface contaminant transport. They
argue that state-of-the-science methods may cost more at the outset, but may yield overall benefits in the form of
reduced clean-up costs conpared to conventional methods. The authors make the point that ground water
processes are difficult to understand and to model due to interactions which may not be simple to describe.
Biotransformations in the presence of dissolved oxygen in an aquifer represents an example where research results
may pave the way for reduced clean up costs at many sites. If more detailed data can be obtained about potential
pathways and mechanisms of transport, the state-of-the-art will be advanced along with the potential for less
costly site restoration.
The major physical processes of importance in ground water transport are advection and dispersion (Freeze and
Cherry, 1979). Advection is the transport of a contaminant by the bulk ground water flow. Dispersion is the
spreading of a contaminant front due to molecular diffusion and small scale variations in fluid velocity throughout
the aquifer.
The major chemical processes of interest are adsorption, ion exchange, hydrolysis, and oxidation-reduction
reactions. Adsorption is "the process in which matter is extracted from die solution phase and concentrated on
the surface of the solid material" (Weber, 1972).
Dispersion--
Dispersion, the spreading of a contaminant front as it moves in the ground water, is an area of particular
controversy at this time. The dispersion process can be most easily described as consisting of three components:
1) molecular diffusion resulting from Brownian motion of individual molecules; 2) hydrodynamic dispersion
resulting from variations in interstitial pore velocities; and 3) macrodispersion resulting from structural variations
in hydraulic conductivity and, consequently, in velocity. Differences in permeability between layers can result in
different ground water velocities and large variations in solute concentration. When an aquifer is monitored using
a fully screened well, ground water from different layers is mixed, resulting in a smoothing of the apparent solute
67
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breakthrough curve. This smoothing can result in very large apparent dispersion when matched against simple
two dimensional solute transport models.
The effects of molecular diffusion, hydrodynamic dispersion and macro-dispersion are frequently combined to
form a dispersivity tensor which in some cases can be reduced to three main components: longitudinal dispersivity
(aj), transverse dispersivity (at), and vertical dispersivity (av). Dispersion coefficients (D) used in the advection
dispersion equation are found by
D = avm (2-10)
where
v is the resultant velocity scalar
m is assumed equal to 1.0
The physics and mathematics necessary to describe molecular diffusion and hydrodynamic dispersion are well
established. Bear (1979) and Fried (1975) provide comprehensive experimental and theoretical reviews of these
processes. The significance of the third component, macrodispersion, is the subject of much debate. Anderson
(1981) summarizes much of the current research on the nature and significance of macrodispersion. At present
there appears to be two dominant approaches.
1. Macrodispersion occurs due to random variations in permeability which can never be adequately
characterized; consequently, the only reasonable method is to employ a stochastic procedure for
describing the average movement of a solute; or
2. The apparent spreading of many solute fronts is due to variations in permeability which are complex but
measureable; consequently, more effort should be expended towards measuring the actual permeability
distributions and using these as input for deterministic simulations.
No work has yet been focused on the effect of varying aquifer parameters on solute transport as it relates to
biorestoration. At present, the predictive accuracy of biorestoration modeling is severely limited by uncertainties
in solute transport simulations. When simulating the transport of a contaminant and oxygen or other nutrient, the
most commonly used numerical models will predict significant mixing between the contaminants and oxygen and
high rates of biodegradation. In real aquifers, contaminants may be trapped in a few areas of low permeability
while the oxygen or other nutrients are forced through the high permeability zones. In this situation, little mixing
of the contaminant and oxygen will occur, and consequently, little biodegradation. Until solute transport models
are developed which can adequately describe the complexities of subsurface flow, the accuracy in which
biorestoration can be simulated will be limited.
Chemical Processes--
The major chemical processes which affect the transport of organic contaminants in ground water are adsorption
and hydrolysis.
Adsorption is a surface process in which a compound "sticks" to the solid aquifer material. In the case of
neutral, nonpolar organics, this stickiness is due to the much higher affinity of the compound for other organics
attached to the soil than for the polar water phase. In the case of polar molecules, adsorption may be due to
dipole:dipole forces. The attraction due to mis mechanism is typically much weaker than that for hydrophobic
compounds.
Naturally occurring organic material in aquifers is commonly present as a humic-kerogen film over the clay
particles. This organic material may originate from humic or fulvic acids deposited with the original sediment or
from infiltrating rain water. Organic material is effectively preserved in tight clays where diffusion of oxygen is
limited and the redox potential is low.
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Hydrolysis reactions can have a major impact on the mobility of organic compounds in aquifers. These reactions
are typically pH dependent and catalyzed by metal ions. The pH of most solutions will approach equilibrium
shortly after entering an aquifer. In this case, hydrolysis reactions can be modeled as a simple first order decay
process.
When attempting to develop models for simulating the adsorption of contaminants in ground water, many early
investigators assumed that at low concentrations, contaminants move independently of other solutes, the
reaction kinetics are fast relative to ground water flow and the natural reactants are uniformly distributed
throughout the aquifer. These assumptions allow the reactions to be analyzed using the equilibrium isotherm
approach. Under this approach the variation in adsorbed contaminant concentration is described by an adsorption
isotherm:
S = f(C) (2-11)
where C = concentration of the contaminant in solution and S is the concentration in the nonmobile solid phase.
This relationhip can be incorporated into the advection dispersion equation by considering the loss of solute to
adsorption.
If S is a linear function of C then the effect of adsorption can be replaced by a constant retardation factor (R). In
this special case the adsorbed contaminant will move according to Equation 2-2 with an effective velocity, v',
where v' = v/R and an effective dispersion coefficient, D' = D/R. A common method of calculating R is by the
relation (Freeze and Cherry, 1979).
R = 1 + n Kdfp (2-12>
where n is the aquifer porosity, p the bulk density, and Kd the partition coefficient in grams of contaminant
adsorbed per gram aquifer.
The use of a retardation factor depends on the following assumptions:
1. Adsorption can be described by a linear relationship between solute and solid phase concentration.
2. The reaction kinetics are fast relative to ground water flow.
3. Natural reactants attached to the aquifer material can be assumed uniformly distributed in space.
4. Contaminant transport is independent of other liquid phase organics.
5. Mathematical Models of Subsurface Biorestoration
Equation 2-2 with terms included for biodegradation can be solved to obtain the concentration of the contaminant
in space and time by both analytical and numerical methods. Analytical solutions generally require less effort for
the model user to employ, but may also require simplification of the aquifer conditions and biodegradation
processes. For uniform flow in an aquifer in which biodegradation may be approximated as a first order decay,
solute concentrations in space and time can be calculated using the one dimensional solution of Ogata and Banks
(1961) or the two dimensional solution of Wilson and Miller (1978). Approximating biodegradation as a first
order decay would be appropriate when the microbial concentration is constant, growth is only dependent on the
contaminant concentration, and the contaminant concentration is significantly less than the half saturation
constant (K). Simkins and Alexander (1984) provide useful guidelines for determining when consumption of
substrate may be approximated as a first order decay.
Numerical solution of Equation 2-2 allows the user much more flexibility in specifying aquifer geometry and
biodegradation kinetics. The most common mathematical formulations for approximating the solute transport
equation are finite differences, finite elements, and the methods of characteristics.
Finite difference models have been developed for a variety of field situations including saturated and unsaturatd
flow, and for transient and constant pollutant sources. Finite difference methods operate by dividing space into
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rectilinear cells along the coordinate axes. Homogeneous values within each cell are represented by values at a
single node. Partial differentials can then be approximated by finite differences and the resulting set of equations
solved by iteration (Mercer and Faust, 1981; Prickett, 1975). Approximating the differentials by a difference
requires that the remaining terms of the Taylor's expansion be dropped resulting in a truncation error and
signficant spreading of the simulated contaminant front. This spreading has been termed numerical dispersion and
can often mask the actual physical dispersion process (Anderson, 1979).
The finite element method also operates by breaking the flow field into elements, but in this case the elements
may vary in size and shape. In the case of a triangular element, the geometry would be described by the three
corner nodes where heads and concentrations are computed. The head or concentration within an element is
allowed to vary in proportion to the distance to these nodes. Complex interpolating schemes are sometimes used
to predict parameter values accurately within an element and thereby reduce the truncation errors common in
finite difference procedures. Some numerical dispersion may still occur but is usually much less significant. The
use of variable size and shape elements also allows greater flexibility in the analysis of moving boundary
problems which occur when there is a moving water table or when contaminant and flow transport must be
analyzed as a coupled problem. A disadvantage of the finite element method is the greater mathematical
complexity and generally higher computing costs (Finder and Gray, 1977; Wang and Anderson, 1982).
The method of characteristics (MOC) is most useful where solute transport is dominated by convective transport
One of the most commonly used models employs a procedure where idealized particles are tracked through the
flow field (Konikow and Bredehoeft, 1978). In step one, a particle and associated mass of contaminant is
translated a certain distance according to the flow velocity. The second step adds on the effect of longitudinal and
transverse dispersion and sources and sinks for the contaminant.
All of these techniques can be used to simulate in situ biorestoration under certain circumstances although no
single procedure will be applicable to every situation. Only very limited work has been done on simulating the
simultaneous effects of advection, dispersion, and chemical and biological processes. In the following section, the
few studies that have been performed are reviewed and the potential weakness of each technique discussed.
Kosson et al. (1985) employ a simple one-dimensional finite difference solution to simulate the movement of
hazardous industrial wastewaters through an acclimated soil column. Adsorption is assumed to be linear and is
described by a retardation factor. A portion of the influent wastewater is assumed nondegradable. Biodegradation
of the remainder is simulated as a first order decay. Experimental data are also provided by a field scale column
used to study the degradation process. The model adequately matches experimental data from the later portion of
the column biodegradation test when an acclimated microbial population had developed. Agreement between
model and experimental results is not as good during the earlier part of the test before the microbial population
had reached steady state.
Angelakis and Rolston (1985) present a mathematical model for simulating the movement of insoluble
(paniculate) and soluble organic carbon through the unsaturated soil profile. Transformation from insoluble to
soluble and finally to carbon dioxide is assumed to follow first order kinetics. Transport of carbon dioxide is by
gaseous diffusion. Simulation results are obtained from analytical and numerical solutions. These results
compare favorably with experimental data from a series of column tests performed using primary wastewater
effluent. Insoluble and soluble organic carbon distributions were adequately matched. A variable gaseous
diffusion coefficient was required to match the observed carbon dioxide distribution.
Baehr and Corapcioglu (1985) present a one dimensional model for simulating gasoline transport in the
unsaturated zone which includes transport by air, water, and free hydrocarbon phases. The hydrocarbon is
assumed to be composed of n components of differing solubility and volatility. Exchange between the air, water,
hydrocarbon and adsorbed phases is assumed to be rapid and described by equilibrium partition coefficients.
Biodegradation of the hydrocarbon is limited by the availability of oxygen which can enter the soil dissolved in
the water phase or by gaseous diffusion. Microorganism growth was not simulated directly since biodegradation
was assumed to be rapid relative to mass transport and to be limited by the availability of oxygen. The equations
are solved numerically using a finite difference procedure. Model simulations indicated that the rate of
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biodegradation was very sensitive to the diffusive properties of the soil. No experimental data are presented to
test the model predictions.
Sykes et al. (1982) simulate the anaerobic degradation of a landfill leachate plume in the saturated zone at the
Canadian Forces Base in Borden, Ontario. Microbial growth, decay and substrate utilization are simulated using
Monod kinetics. When substrate concentrations are significantly below the half saturation constant and microbial
populations are close to steady state, the biodegradation kinetics are reduced to a first order decay. The nonlinear
equations are generated using a Galerkin finite element approximation and solved using a Newton Raphson
iteration procedure. Model simulations indicate that the majority of the degradable organics can be expected to be
removed within a few meters of the landfill. This finding was confirmed in field studies at the site. Sensitivity
analyses performed using a one dimensional solution indicated that under certain circumstances, pulses of
organics could escape from the landfill before a significant microbial population has developed.
Molz et al. (1986) present a numerical model for simulating substrate and oxygen transport and use by attached
microorganisms. The microbial population is assumed to be immobile and present in microcolonies of an average
thickness and radius. Transport into the microcolonies of oxygen and substrate is limited by diffusion through a
stagnant layer adjacent to the microcolony. Microbial growth and consumption of oxygen and substrate within the
microcolony is described by Monod kinetics. A one dimensional solution is obtained numerically using an
Eulerian-Lagrangian finite element solution. The numerical model will be then used to simulate the transport and
biodegradation of substrate and oxygen in a laboratory column. The simulation results indicate that degradation is
most rapid near the column inlet. The initial microbial population has a significant effect on the simulated
breakthrough at the beginning of the simulation but has little effect on the steady state substrate distribution.
Large substrate loadings at the column inlet quickly exceed the available oxygen supply resulting in anaerobic
conditions throughout the majority of the column. Laboratory testing of the simulation model is planned.
Borden and Bedient (1986) present a numerical model of oxygen limited biodegradation of hydrocarbons in the
saturated zone. Numerical solutions are obtained by approximating one dimensional flow as a series of
completely mixed reactors and two dimensional flow using an explicit finite difference solution corrected for
numerical dispersion. One dimensional model simulations indicate that biodegradation will be very rapid near the
contaminant source when oxygen is present. When no oxygen is present at the source, biodegradation will be
slow and limited by the transport of oxygen into the contaminant plume. Two dimensional simulations indicate
that horizontal and vertical mixing are the major sources of oxygen to the contaminant plume and control the
biodegradation process. When adsorption of the hydrocarbon to the aquifer is significant, advective fluxes of
oxygen into the plume and resulting biodegradation is also significant. Sensitivity analyses with the model
suggest that for many aquifers, the reaction between oxygen and hydrocarbon may be approximated as an
instantaneous reaction since oxygen transport is rate limiting. Borden et al. (1986) employ these results to modify
the USGS Solute Transport Model (Konikow and Bredehoeft, 1978) to simulate oxygen-limited biodegradation of
creosote wastes at a Superfund site. The model gave an adequate description of the observed oxygen and
hydrocarbon distributions in the shallow aquifer at the site and was used to study various remedial actions
including no action and removal of the contaminant source.
Dawson et al. (1986) modify a petroleum reservoir code to simulate enhanced in situ biorestoration using the
equations presented by Borden and Bedient (1986). Advective and dispersive transport is calculated using a finite
element-modified method of characteristics solution which allows a large time step and strongly advection
dominated flow. Because the rates of biodegradation can be very high relative to transport, a time splitting
scheme is employed where the microbial kinetic terms are solved separately using an implicit solution with a
much smaller time step. This model is then employed to simulate enhanced in situ biorestoration by the injection
of oxygen and production of contaminated water by a five-spot pattern. Simulations are performed for a variety of
conditions including uniform and random permeabilities and variable adsorption.
6. Model Use and Limitations
The current technology for simulating subsurface biorestoration is still in its infancy. Some progress has been
made in developing kinetic descriptions of the biodegradation process and combining these with available solute
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transport models. Unfortunately, little reliable field data has been available to rigorously test these models.
Considerable uncertainty exists over the importance of simulating transport into biofilms or microcolonies. Also,
the effects of variations in aquifer parameters on the efficiency of biorestoration is unknown. At present, the
technology is not available to quantitatively predict the efficiency of enhanced biorestoration but significant
advances are being made in our ability to describe the process.
7. Conclusions
Of the available biological aquifer remediation techniques, the most effective methods are enhancement of the
native population and withdrawal and treatment by various wastewater treatment processes (Lee and Ward, 1984).
Before any aquifer remediation technique can be implemented, a thorough understanding of the hydrogeology and
contamination problems of the site must be obtained and used to design the treatment sytem (Lee et al., 1986).
Costs for treatment may range from tens of thousands up to tens of millions, depending upon the extent and nature
of the contaminants, the nature of the site, and the desired cleanup levels.
Addition of oxygen, nitrogn, phosphorus, and trace minerals stimulates the acclimated indigeneous microbial
population to aerobically degrade the contaminants. In situ biorestoration has been chiefly used to treat gasoline
contaminated aquifers, but also has been employed with ethylene glycol and solvents including acetone,
tetrahydrofuran, methylene chloride, n-butanol, dimethyl aniline, and isopropanol. Biorestoration effectiveness
will be affected by toxic levels of organics and heavy metals. In general, in situ bioreclamation has been effective
in reducing the quantity of the contaminants, but not in completely eliminating them. The treatment moves with
the plume allowing treatment of trapped or sorbed contaminants or, by using soil flushing or an infiltration
gallery, in sjtu microbial treatment can reach areas that are not accessible by other techniques. Biorestoration has
been used in a number of aquifers, but may be of limited usefulness in those with low permeabilities. Undesirable
metabolic and inorganic nutrients may escape from the treatment zone and affect ground water or surface water
quality. Alternative oxygen sources such as ozone, hydrogen peroxide, pure oxygen, and air flooding or soil
venting may speed the removal of the organic contaminants, but their impact on the microbial population and the
geochemistry of the site is not fully understood. Innovative processes such as treatment beds or land treatment
can be used in some situations. In the presence of an acclimated microbial population, many aquifers will be
anaerobic because the microorganisms will have depleted the dissolved oxygen. It will be possible to use
anaerobic degradation to remove contaminants, although the technology for this treatment has not yet been
developed. Reducing the interfacial tension between the hydrocarbon and ground water with surfactants,
dispersants, or emulsifiers will mobilize the contaminants and may make them available for microbial
degradation. Combinations of in situ biorestoration treatment with other chemical, physical, or biological
treatment proceses have been successfully utilized in aquifer remediation.
Treatment by biological wastewater processes is a proven technology. The biological processes include activated
sludge, lagoons, waste stabilization ponds, fluidized bed reactors, trickling filters, rotating biological discs, and
sequencing batch reactor. All of these processes are dependent upon extraction of the contaminated ground water
from the subsurface. Combinations of conventional wastewater treatment processes and other water treatment
processes have also been successful.
Alteration of the subsurface microbial community has a great deal of potential to allow degradation of recalcitrant
compounds in the subsurface. The added organisms are selected by enrichment culturing or genetic manipulation.
However, introduction of non-native microorganisms may be limited by movement of the organisms through the
subsurface, survival of the organisms, and accessibility of the organic contaminants. Addition of an acclimated
population may be more successful when combined with wastewater treatment processes where the environment
can be more closely regulated. Although the aquifer remedial actions that have used a microbial seed have not
conclusively shown that the added organisms were responsible for removal of the contaminant, the concentrations
of the contaminants were reduced. Alteration of the environment to promote the activity of a particular
component of the microbial community is another promising technology. Field tests and further research are
currently underway for this technology. The environment is altered to promote the growth of organisms that co-
oxidize halogenated aliphatics when supported on gaseous hydrocarbons such as methane, propane, or natural gas.
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Techniques for simulating the subsurface biorestoration process are under development, but little reliable field
data has been generated that can be applied to these models. Some of the major considerations in simulating
transport and biodegradation of organic contaminants in the subsurface are poorly understood; these include the
importance of transport of organics to the bacteria and the variability in aquifer parameters.
III. Institutional Limitations on Ground Water Pollution
Control
The development and application of technologies for the prevention and regulation of leaking underground storage
tanks is a complex and interdisciplinary science, as is the abatement and cleanup of the ground water
contamination threats such leaks present. This science is further complicated, however, by the institutional
limitations that control the rate at which technology can advance and its use in the development of regulations.
Institutional limitations may be the factor in determining how the technology will be applied to leaking
underground storage tank regulation and remediation. These institutional limitations that control the rate at which
technology can advance and its use in the development of regulations. Institutional limitations may be the factor
in determining how the technology will be applied to leaking underground storage tank regulation and
remediation. These institutional limitations include: (1) the scientific understanding of the nature of leaking
underground storage tanks and released products; (2) public opinion; (3) business community attitudes; (4)
environmental interest group concerns; and (5) governmental acceptance, use, and implementation of the
remedial technology.
A. Scientific Understanding of the Nature of Released Products from Leaking
Underground Storage Tanks
Although continuously researched and tested, the scientific understanding of the nature of a released product and
its behavior in the subsurface, as well as the causes of leaks in underground storage tanks, is still rather limited.
These subjects are being researched in laboratory microcosms and through the use of models, but field conditions
continue to produce situations where only limited scientific understanding is available for the development of
remedial technologies. These limitations can be described within the following categories: (1) the physical and
chemical nature of petroleum products in the subsurface; (2) the nature and occurrence of leaks in underground
storage tanks; and (3) hydrogeologic complications associated with the migration of contaminant plumes.
1. Physical and Chemical Nature of Petroleum Products
Petroleum products commonly stored in underground storage tanks are labeled with familiar names such as
"regular", "unleaded", and "super-unleaded" gasoline, diesel, and kerosene. While the names infer a single
compound, they are misleading in that they actually refer to a very complex mixture of chemicals which can
include a variety of hydrocarbons and additives. These multiple components have individual physical and
chemical characteristics which can interact and react independently when introduced into the subsurface from a
leaking tank. Petroleum products from three separate phases during their migration through soil media and in
their contact with ground water: (1) free product phase; (2) dissolved phase; and (3) vapor phase.
The free product phase describes the body of the product which retains its basic identity and composition in the
subsurface. The main portion of the contaminant plume meets this description, but so does the residual product
which is retained in the subsurface media as the plume is transported away from the leaking underground storage
tank. The dissolved phase describes that portion of the released product which goes into solution with contacted
ground water. Because the product is a mixture of chemicals, compounds react and dissolve with the ground
water based on their individual properties and their individual equilibrium reactions, as well as through the effects
of the interacting reactions and combined equilibrium conditions.
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volatilize from both the free (and residual) product phase and the dissolved phase with concentrations based on
both product quantity and individual and combined equilibrium rates.
The three phases also include the portion of the residual product which is inundated as the ground water level
fluctuates with rising and falling water tables. A rising water table may dissolve a portion of the residual product
as well as floating it to reform free product. A falling water table may cause the reformed free product to again
adhere to the subsurface media as a residue. The vapor phase may be evident during this fluctuating process.
The primary scientific limitation involving the nature of the leaked product is associated with understanding,
analyzing, and researching interactions of the multitude of compounds present in the petroleum product with the
multitude of existing subsurface media. Researching the rates and effects of partitioning, synergistic reactions,
volatilization, dissolution, and adsorption for this multitude of compounds is very time consuming and has only
fairly recently been intensively undertaken. Analytical characterization of these compounds can be expensive and
time consuming, and it is very dependent on current developments in analytical techniques and equipment.
In addition, some components of petroleum products are known to degrade in the subsurface through microbial
processes. The study of these processes is critical to the understanding of the fate of petroleum products in the
subsurface.
2. Nature of Occurrence of Leaks
Improper installation, corrosion damage, external damage, internal punctures, and rusting of tank bottoms are
commonly documented causes of underground tank leakage. While the nature of underground storage tank
failures and leaks can be easily determined and are well documented, the actual mechanisms for tank integrity
problems are often complex. The physical and chemical processes involved identifying the combined forces
which result in a tank collapse or failure are complicated and often do not truly define the cause of a failure.
Improper tank installation or external human activities can cause tank leakage. Undetected non-uniform backfill
around a tank, such as a single sharp rock, a piece of metal, and/or irregularly packed backfill, can lead to tank
failure. A single lump of clay or a stray current from adjacent businesses, transportation systems, or utilities can
produce unexpected and rapid corrosion of underground storage tank system components.
Improper maintenance or system abuse can create conditions which result in tank and piping fsilure or system
deterioration. "Spearing" a tank bottom while checking the tank product level, or bouncing the level indicator
device off the tank bottom by inexperienced and unsuspecting operators, can result in immediate tank failure or
the weakening of the tank material beneath the fill spout, the lack of water detection in operating tanks and its
frequent removal can result in tank bottom corrosion and possible tank failure. Improper maintenance of
protective cover pads over the underground storage tank system can result in crushed lines or a collapsed tank due
to heavy vehicular traffic over deteriorated areas. These relatively minor and often unseen conditions can easily
undermine the best constructed underground storage tank system.
The nature and occurrence of leaks and spills from underground storage tanks is also more complex than is
generally believed. Current discussions regarding regulatory requirements to detect leaks often center on the
installation of monitoring wells around underground storage tank systems. These wells may even be required to
be equipped with hydrocarbon detectors in lieu of periodic sampling. However, what will the detection of
hydrocarbons in the monitoring wells tell the underground storage tank system owner or operator? The
contamination could come from surface spills at the service islands, transfer spills, overfill spills, or even spills on
adjacent or distant properties. The implication that contamination in the wells would clearly indicate a leak in the
underground storage tank system could cost the owner or operator a great deal of money for integrity tests, tank or
piping removal, additional bore holes or monitoring well drilling, or legal fees. The determination of the source
of subsurface hydrocarbons is a critical factor for the implementation of any underground storage tank regulations
or the development of remedial technology.
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The location of the deterioration in the tank can often produce leaked product or sometimes not be a constant
environmental threat. If a deterioration occurs in the bottom of the tank or below the lowest maintained product
level, a continuous leak can occur. If the deterioration is slight, a leak could occur only when the hydraulic head
is great enough to drive the product out of the tank. If a deterioration exists below the highest maintained
inventory but above the lowest level, an intermittent leak will occur whose magnitude will be related to the
volume of product above the leak. The deterioration may even exist above the highest maintained level and a leak
may never occur.
The water table can affect the rate of a leak as well as the existence of a leak. If a tank is inundated by ground
water, the flow of product from the tank versus the flow of ground water into the tank is controlled by the pressure
difference of each liquid's hydraulic head. If the tank product has a greater head, product can leak out. If the
ground water has a greater head, water will enter the tank.
The predictability of tank leaks, and extent of product migration into the subsurface environment, and their
remediation are limited to knowing when a deterioration in a tank is acting as a leak source. The prediction of a
constant leak from the bottom of a tank is relatively simple and can be simulated with a number of computer
models. A sporadic leak is much more difficult to simulate. Such a leak may never reach a magnitude that can
even be detected, or it can produce individual spills of varying volumes at widely varying times dependent on
inventory volumes.
Pipeline leaks can also be very erratic and hard to predict and evaluate. The types of pumps used in transferring
product from the tank to the dispenser directly affect the occurrence of leaks from a pipeline system. A positive
head pump pushes the product into the pipeline. If a pipeline leak occurs, the positive pressure on the pipeline
forces the product out of the system. A negative head pump pulls the product from the tank to the dispenser. If a
pipeline leak occurs, the vacuum created on the pipeline draws air into the system. Although little or no product
is lost while the negative pump operates, the pipelines still contain product while the system is passive, as with the
positive head system, and product can leak from the system. This erratic loss of product has the same limiting
impact on the predictability and evaluation of piping leaks as do the occurrence of leaks from tanks.
3. Hydrogeologic Complications
Once a tank product is introduced into the subsurface from a leak, a detailed understanding of the physical and
chemical nature of the contaminant and its interaction with the subsurface, and the local hydrogeologic formations
and features through which it migrates, is essential. Any technology developed for the remediation of
underground storage tanks must take into account the physical characteristics of the subsurface which dictate the
migration path of the contaminants. Unfortunately, the subsurface is not always a homogenous formation with
predictable, and reproducible effects on contaminant plume movement. The complexities of lens formations,
karst formations, fracture formations, irregular confining beds, perched aquifers, recharge zones, etc., are rather
well documented and clearly represent a limiting factor in the development of remedial technologies for leaking
underground storage tanks.
Additional limiting factors in the development of remedial technologies are unnatural features present around
underground storage tank systems which affect the fate of leaks from these systems. These factors include: (1)
the fill material surrounding tanks and pipelines; and (2) underground utility and service lines and their associated
backfill. Backfill material varies around underground storage tank systems from pea-sized gravel to original
excavation material, from similar backfill around the entire underground system to one type of backfill around the
tanks of a system with a different type around the pipelines, from old tanks or piping having original excavation
material backfill to new or replacement tanks or lines having imported backfill, etc. The older the underground
storage tank system, the more variable the backfills and inconsistencies which can be expected around each
component of that system.
If a tank has a backfill that is more permeable than the subsurface surrounding the excavation, leaked product can
pool within the excavation and migrate into the environment following a path of least resistance. The product
might overflow the excavation and migrate along the land surface or along road foundations or into adjacent
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building backfills. The product might pool until it migrates vertically through the underlying strata. The product
could even pool within the excavation and act as a hydraulic seal preventing additional product from leaking out
of the tank.
If a tank has a backfill that is less permeable than the surrounding material, the migration of the leaked product
can be, at first, held intimately to the tank's external surface until the backfill becomes saturated and the product
breaks through to the surrounding subsurface material. Once the product breaks through, its migration is
controlled by the characteristics of this subsurface material.
Once product migrates away from the tank excavation, it can be intercepted by utility and service lines such as
sanitary sewer lines, storm sewer lines, water lines, telephone service lines, natural gas lines, etc. The migrating
product can follow these lines using surface tension or, if the backfill of the lines is more permeable than the
natural substrate, can preferentially follow a line with a large portion of the contaminant plume becoming
involved. As the product migrates along the lines, it can degrade seals or the line material itself, and can cause
extensive damage or enter the line system. If the line is a sanitary sewer line, the product can appear in toilets or
sinks a great distance away from the leak source without affecting hook-ups between the source and the surfaced
product. If the line is a water line, the contamination can be widespread with a variety of effects. Telephone
service lines and manholes can become storage structures for leaked product, not being discovered until service
problems arise and the service personnel attempt to enter the manhole.
These unnatural features can divert the migrating plume along very erratic and unexpected routes. They present
themselves, in certain situations, as radical limiting factors in the development of remedial technologies for
underground storage tanks.
B. Public Opinion
A very large portion of the public owns automobiles and relies heavily on petroleum products. Therefore,
awareness of the problem of service station leaks and the forthcoming regulations is fairly common. Public
opinion towards the development and implementation of underground storage tank programs can be grouped
under three headings: (1) persons not generally concerned with, but who will openly accept the forthcoming
regulations due to either apathy, ignorance, or having a clear understanding of the need for such regulations; (2)
persons concerned about increasing government control over and interference with private industry; and (3)
persons who have been affected or who fear they will be affected by leaking underground storage tank incidents.
A portion of the public will accept forthcoming underground storage tank regulations because they feel that they
will not be very affected by them. They will be willing to accept increased costs for petroleum which may occur
with the new requirements, or they may not yet be aware that higher costs may result. They may have heard of
the problems with leaking underground storage tanks and realize the need to remedy the potential threat. The
apathetic portion of the public may not realize the potential commercial impacts of such regulations, or they may
feel that such impacts are acceptable. A large group may well understand the problems related to leaking
underground storage tanks and will openly accept any programs and associated impacts in order to control the
potential and existing threat.
Another portion of the public is very concerned over increased governmental control over private industry, thus
they consider attempts to regulate this problem as interference. They feel that the industry can take care of itself.
This group would include a large portion of the small station owners and related businesses who may not be able
to survive stringent regulations. They look at past government involvement with farmers, the past oil embargo,
and the history and related press involving price controls, wheat subsidies, etc., and fear government involvement
They especially fear the potential impact on small businesses and the common fear of the "rich getting richer and
the poor getting poorer". The recent economic condition of the country fuels these fears as they see that it gets
harder and harder for small businesses to survive. The concerns of this portion of the public are very real and
must be considered as the technology is developed for the remediation of leaking underground storage tanks, and
the forthcoming underground storage tank programs and associated regulations.
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A growing portion of the public, although thankfully still a minority, are those who have been affected by leaking
underground storage tank incidents or know of persons who have been affected. These people have lost then-
water wells, have had their water supplies contaminated, have had flammable fumes enter their homes or
businesses, or have lost property due to explosions or extensive contamination. Others have lost their businesses
due to exorbitant clean-up costs for leaks they are legally responsible for, have paid exorbitant legal fees when
they have been sued for damages, pay higher and higher insurance, have had to pay for new water wells, or have
lost farms or homes because Uiey could not find an uncontaminated water supply after theirs was lost due to
leaking underground storage tanks. Still others feel they have been given the runaround by the government to
which they have gone for help. They feel exasperated because no one is able to solve the contamination problem
which has crept beneath them. They feel helpless as their health worsens and their lives fall apart as a result of
drinking undetected contaminated water, or as they breathe fumes from leaked products. These persons will
probably have the greatest public voice in the development of leaking underground storage tank remedial
technologies and underground storage tank programs.
C. Business Community Attitudes
The business community and their general attitudes towards the development and implementation of underground
storage tank programs can be grouped into four categories: (1) underground storage tank owners and operators,
including retail businesses and companies which maintain their own underground storage tank systems not for
retail sales use; (2) insurance companies; (3) petroleum product sales support and maintenance companies; and (4)
the general business community.
1. Underground Storage Tank Owners and Operators
Underground storage tank owners and operators will be the group most affected by underground storage tank
regulations. This group includes refineries, chemical plants, and retail stations as well as other businesses which
maintain underground storage tanks for vehicle fleets such as bus companies, ice companies, automobile
dealerships, etc. Even local, state, and federal government entities maintain underground storage tanks for vehicle
fleets. The concerns of the underground storage tank owners and operator fall into three main areas: (1) is their
system currently leaking and how can they adapt to new requirements; (2) can they financially survive such
regulations; and (3) can the government develop regulations that solve the problem without going to extremes?
When asked, most owners and operators will emphatically state that their underground storage tank system is not
leaking. An alarmingly large number of small business operators, however, do not keep adequate inventory
records and are often not aware of the signs that their system may not be tight They know that if the system is
leaking, the financial outlay to correct the problem will probably put them out of business. Without adequate
funds, survival of the business relies on hoping the system is tight. Larger companies that have adequate funds for
system and component testing, repair, and replacement will quickly provide proof of system tightness or take
appropriate action to remedy a leaking system.
Small businesses fear the implementation of underground storage tank programs because current economic
conditions have severely limited their profits. If required, these businesses could not afford to provide proof that
their system is tight, and would probably go out of business long before they were required to retrofit their system
with leak protection or detection. These businesses realize that regulations are being developed, realize that they
may not survive, but will continue to operate while they can. Owners of these businesses will probably oppose the
implementation of regulations but realize their survival is short-lived.
Larger, privately owned stations can usually afford to adequately maintain their underground storage tank
systems, keep close watch of their inventory, investigate possible signs of leaks, and make necessary repairs,
replacements, and cleanups when necessary. These businesses are more concerned over how strict the regulations
will be, and how they will have to alter their operation to survive. They realize that it takes money to maintain an
underground storage tank and a business and they are, hopefully, making enough profit to adapt to forthcoming
underground storage tank program requirements.
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Currently, corporation-owned stations are generally already implementing programs that they perceive an
underground storage tank program might include. They realize that by the time the regulations are implemented,
inflation will cause the installation of required leak prevention and detection equipment to cost a great deal more
than they would now. They also feel that, because of their size, own staff expertise, and potential lobbying
capabilities, they can convince regulators that their systems are tight and well-protected. They may however, fear
for their survival should the new regulations be more stringent than they have anticipated.
Also in response to the forthcoming regulations, major oil companies are cutting back on the number of stations
they operate. Many stations are being sold. The buyers of these stations are very aware of the potential problems
with such stations, and are requiring site contamination evaluations prior to the sale as a condition of the sales
agreement. The sale of these stations lessens the chances for pollution liability over time for the seller by getting
rid of sites which may be considered high potential pollution risks. Buyers must realize this and be willing to
make the necessary improvements to abate risk once the site has been shown to be clean or has been cleaned up.
A major concern of all underground storage tank owners is the ability of the government to develop reasonable
regulations in its rush to regulate underground storage tanks and get control of the leaking underground storage
tank problem. Large businesses are concerned about their international competitiveness if regulation becomes
excessive. Small businesses are very concerned that the developing regulations may not consider the smaller,
privately owned stations and small oil companies. Small businesses rely on petroleum marketing associations to
actively lobby state and federal governments in an effort to assure reasonable regulation development Large oil
companies, in addition to their lobbying efforts, are sponsoring or co-sponsoring, with consulting firms and related
associations, educational seminars on the prevention of leaking underground storage tanks, maintenance of
underground storage tanks, and existing state and local regulation of underground storage tanks. These seminars
can help educate government employees so that reasonable and intelligent regulations can be developed.
2. Insurance Companies
An additional factor that all underground storage tank owners are concerned about is having adequate insurance
coverage in case a leaking underground storage tank incident occurs. The concern is greater in smaller businesses,
but still exists even in the large corporations. Insurance companies are also very concerned with this problem. In
recent years the number of leaking underground storage tank incidents has dramatically
increased. Technology has developed to analyze the impacts of the associated contamination, but the costs of
incident evaluation, abatement, and remediation can be extensive. As insurance companies have had to pay out
more and more for these episodes, policies are now being cancelled, greater restrictions are being placed on
policies that are issued, and the cost of insurance is soaring.
In an effort to control costs, an increasing number of insurance companies are contracting their own
environmental consultants to cleanup leaking underground storage tank incidents experienced by policy holders.
The consultants are also contracted to drill monitoring wells at most stations of policy holders to detect leaks
before the cost of handling a pollution incident become even more expensive. This concern for cost control
includes a concern that governmental bureaucracies can develop reasonable and adequate regulations which can
help stem the number of leak events and, thereby, lessen the number and related expenses of policy claims.
3. Petroleum Product Sales Support and Maintenance Companies
Another business group which can be directly affected by underground storage tank programs is the petroleum
product sales support and maintenance companies. If the forthcoming regulations force stations out of business,
these companies could realize a major loss of business. However, these businesses could realize an increase in
business by dealing in leak prevention and detection equipment. The economic hard times for retail stations can
also increase the wholesale business of service-related products as stations sell additional products to increase
profits.
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This group of businesses often deals in state-of-the-art leak prevention and detection products which can result in
high business risks. Not all of the products currently on the market have been adequately tested, proven to
perform their advertised claims, been proven to work adequately through time, or been tested as having a potential
for showing false readings. If regulations are implemented which restrict the use of uncertified equipment or
require certain minimum standards which stockpiled detection or prevention equipment do not meet, wholesale or
service businesses could be as adversely affected as underground storage tank owners. Currently, governmental
bureaucracies provide little or no guidance or control for this type of equipment. As a result, these support
businesses, which operate on a demand basis, have great interest in and concern over the development of
underground storage tank regulations.
4. General Business Community
In addition to the petroleum-related businesses, the general business community also has concerns over the
development of underground storage tank regulations. Private business is wary of government involvement in
private industry. They often resent additional governmental control over problems which could be handled by an
industry itself. They see increased costs resulting from such involvement.
D. Environmental Interest Groups
Environmental interest groups are very aware of the existing and potential problems with leaking underground
storage tanks. They understand the toxic and carcinogenic nature of petroleum products and the impact that these
compounds can have in the environment. They see the immediate need for the development and implementation
of underground storage tank programs, but doubt the government's ability to perform the task.
These groups are concerned that inadequate regulations may be implemented because the government is unaware
of the severity of the problem. They are concerned that the forthcoming regulations may not require enough
safeguards on underground storage tank systems due to lobbying efforts by big business. The final requirements
may also tend to give preferential treatment to big business due to these strong lobbying efforts and political
interactions.
Environmental groups also understand the threat posed by abandoned underground storage tanks, and are
concerned that all of these tanks may not be located or investigated. They are concerned that local and state
regulations may also be inadequate but may still be accepted by the federal government due to bureaucratic and
political reasons.
They are also concerned over the scientific and technological limitations for detecting and evaluating the threat
posed by leaked products. This basic lack of information may result in inadequate limitations for final cleanup of
contamination incidents. They are further concerned that this lack of adequate scientific knowledge, combined
with the finalization of underground storage tank regulations, will not provide the needed safeguards to protect
environmentally sensitive areas which may be affected by leaking underground storage tank contamination
E. Government Agencies
Government agencies are faced with a multitude of problems when both dealing with leaking underground storage
tanks and the development of timely and effective underground storage tank programs. These problems can be
categorized as: (1) unclear jurisdictions, (2) "Who is responsible for cleanup?" (3) "What is an adequate
underground storage tank program?" (4) "How clean is clean?" and (5) general institutional concerns.
1. Unclear Jurisdictions
The effects of leaking underground storage tanks and petroleum product contamination are complex and highly
variable. Environmental pollution, toxic and carcinogenic chemical health threats, and fire and explosion damage
are often direct dangers from petroleum product pollution from leaking underground storage tanks. Product loss,
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from a governmental view, also includes the loss of tax revenues and possible loss of controlled product quality.
State and local governments are often not organized in a fashion which grants the jurisdictional responsibility for
these effects to any single agency. Yet, all of these problems must be addressed during leaking underground
storage tank incidents and the development of underground storage tank programs. Which state or local agency
should receive jurisdiction for an underground storage tank program, and which can organize and handle all of
these various problems?
State governments will be given the lead in underground storage tank programs in most states, with local
ordinances given priority if considered more or equally stringent. Within state government, however, the
organizational structure may not be conducive to easily determining underground storage tank jurisdiction. States
with single environmental or natural resource departments can fairly easily endow these organizations with the
responsibility. Other states which are not so organized may struggle with the determination. Political power plays
and in-fighting for related underground storage tank funds and responsibility could result, especially in
economically-strained states. Response by the government agencies to leaking underground storage tank incidents
and dangers may be delayed because the incidents must be carefully examined to determine whether the current
threat falls under local, state, or federal jurisdiction.
2. Who is Responsible for Cleanup
When a leaking underground storage tank incident becomes evident, the source or the responsible party cannot
always be located. Who is then capable of evaluating the extent of the contamination, abating the effects, and
cleaning up the contamination? The evaluation and cleanup of leaking underground storage tank incidents is a
very costly and time-consuming task.
State and local governments are not always equipped, trained, or financially capable of properly accomplishing
cleanup, or paying to have it done. Yet the incidents exist and their associated dangers and threats to property and
human health can become increasingly evident and widespread. Unless the state or local governments have set up
a tax or fee-based response fund, the agencies and departments, with their various responsibilities, must abate the
threats as they become evident without ever effecting a cure for the incident.
3. What is an Adequate Underground Storage Tank Program?
With the increasing threats posed by leaking underground storage tanks, federal, state, and local governments are
rushing to develop underground storage tank programs to control the problem and related costs of handling
leaking underground storage tank incidents. Unfortunately, the complexities involved with causes and cures for
underground storage tank problems often exceed even the scientific community's understanding, as well as the
government's understanding. How, then, can adequate programs be developed, and who decides what is adequate
and what is not?
The current philosophy appears to be that of watching programs which have been developed by state and local
governments which have been hardest hit by leaking underground storage tank incidents. If these programs prove
themselves in time, certain components of the programs are incorporated into the programs being developed by
other government agencies. These adopted components must, however, be deemed reasonable, functional, and
politically proper by the political climate within each government. Whether this philosophy works must be tested
in time along with the already developed underground storage tank programs.
The problem, however, still exists. What is an adequate underground storage tank program? Are the existing
programs over-regulating underground storage tank systems? Are they underregulated? The results of an
inadequate program can have dramatic effects on the entire business community as well as the local environments.
4. How Clean is Clean?
When a leaking underground storage tank incident occurs and cleanup of the contamination is initiated, when does
the cleanup stop? How clean is clean? This is a continuing question for any pollution remediation undertaking.
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Unfortunately, with the increasing incidents of leaking underground storage tanks, the question arises more
frequently, and its answer can affect nearly every community which has service stations or other types of
underground tanks.
The scientific understanding of the interactions of the components of petroleum products in the subsurface is
unclear. The toxicity and other health impacts of the products are not fully researched. The ability to detect the
various components at potentially health threatening levels is limited and often still in the developmental stage.
At what point will the natural environment be capable of handling residual contamination so that man-initiated
clean-up can stop? These fundamental details must be addressed before a realistic answer can be determined for
the question "How clean is clean?"
Meanwhile, leaking underground storage tank contamination incidents are being cleaned up with no real
guidelines for when remediation is complete. The costs of continued cleanup, monitoring residual contamination,
and unexpected health impacts are expensive and rising. The burden of the expense is absorbed by the business
community, as well as, local, state, and the federal governments. Are the cleanup requirements excessive? Are
they not enough?
5. General Institutional Concerns
The principal criteria for selecting remediation procedures should be the water quality level to which an aquifer
should be restored, and the most economical technology available to reach that quality level. Unfortunately, there
are numerous institutional limitations that sometimes override these criteria in determining if, when, what, and
how remediation will be selected and carried out (Wilson et al., 1986).
Responding to a ground water contamination problem is likely to require compliance with several local, state and
federal pollution control laws and regulations. If the response involves handling hazardous wastes, discharging
substances into the air or surface waters, or the underground injection of wastes, federal pollution laws apply.
These laws do not exempt the activities of federal, state, or local officials or other parties attempting to remediate
contamination events. They apply to generators and responding parties alike, and it is not unusual for these
pollution control laws to conflict. For example, a hazardous waste remediation project may be slowed, altered or
abandoned by the imposition, upon the party undertaking the effort, of elaborate RCRA permit requirements
governing the transport and disposal of hazardous wastes.
In situ remediation procedures may be subject to permitting or other requirements of federal or state underground
injection control programs. Withdrawal and treatment approaches may be subject to regulation under federal or
state air pollution control programs or to pretreatment requirements if contaminated ground water will be
discharged to a municipal wastewater treatment system. Also, pumping from an aquifer may involve a state's
ground water regulations or well construction standards and well spacing requirements as well as interfere with
various competing legal rights to pump ground water.
IV. Research Needs for Optimized Remedial Techniques
Ground water pollution from leaking underground storage tanks represents an issue of recent major attention from
a national perspective. Treatment of contaminated subsurface environments, particularly through the use of in situ
biorestoration, has an even shorter history. Accordingly, as this remediation technology develops, many technical
research needs can be identified. Examples of these needs will be discussed in terms of evaluating the
effectiveness of physical containment techniques, enhancing vadose (unsaturated) zone pollutant removal
techniques, and enhancing microbial population densities.
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A. Evaluation of Effectiveness of Physical Containment Techniques
Physical containment measures include hydraulic barriers, vertical impermeable barriers, such as slurry walls and
grout curtains, and subsurface drains and interceptor trenches. While these measures may be effective in terms of
containing or excluding ground water flow, there are unresolved questions relative to the retention of certain
solutes, both initially and over time. Examples of research questions which need to be answered include:
(1) What is the effectiveness of hydraulic barriers in non-homogeneous subsurface environments?
(2) What is the rate of increase in the permeability of bentonite slurry walls in the presence of certain
organics?
(3) What design approaches can be used to optimize the usage of physical containment techniques (minimize
the costs) in conjunction with pumping, pumping and treating, and/or in situ treatment remediation
measures?
B. Enhanced Vadose Zone Pollutant Removal Techniques
Products released from leaking underground storage tanks can be retained in the vadose zone as a result of both
physical and abiotic processes. Decomposition products may also be retained as a result of biotic degradation and
subsequent abiotic processes. In this context, the vadose zone can be viewed as a "reservoir" which slowly
releases contaminants to the saturated zone over time. Therefore, remediation techniques which would enhance
the removal and/or transport of contaminants from the vadose zone would be desirable. Examples of research
questions which need to be answered include:
(1) What chemical measures can be used to mobilize contaminants held in the vadose zone as a result of
adsorption, ion exchange, precipitation, and/or complexation?
(2) Can man-induced changes in subsurface environmental conditions, such as changes in pH and oxidation-
reduction potential, be used to mobilize contaminants?
(3) Can microbial degradation processes in the vadose zone be optimized through the controlled addition of
nutrients, enzymes, and bacterial seed organisms?
(4) What design approaches and/or laboratory tests can be used to optimize man-controlled hydraulic
flushing of vadose zone pollutants, both with and without recycling of extracted ground water from the
polluted area?
C. Enhancement of Microbial Populations
The current literature indicates that in sjlu biorestoration is dependent upon the indigenous microflora; bacterial
seed organisms are often added in biorestoration treatment schemes but their role in pollutant degradation is yet
undemonstrated. Usually, stimulation of the native microflora results in increases in bacterial populations in the
active biorestoration zone. The increases should coincide with bacterial degradation and may be limited by the
availability of nutrients, lack of optimum environmental conditions for degradation, and/or lack of
microorganisms adapted to the decomposition of leaked organic products. Therefore, increasing microbial
population densities may be necessary in order to optimize the environmental conditions for m situ biorestoration.
Examples of research questions which need to be answered include:
(1) What are the optimum nutrient concentrations for achieving Jn situ biorestoration of different classes of
organic compounds?
(2) What are the optimum environmental conditions (pH, oxidation-reduction potential, micronutrients, etc.)
necessary to achieve in situ biorestoration of different classes of organic compounds?
(3) Can laboratory development of acclimated microorganisms enhance in situ biorestoration, and what
laboratory tests/procedures are necessary to achieve this acclimation?
(4) What design approaches and/or laboratory tests can be used to optimize microbial population densities
for in situ biorestoration?
(5) What are the best methods for achieving in situ mixing of the bacterial populations, nutrients and
micronutrients, and organics?
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(6) What are the optimum (most cost-effective) combinations of ja fiilll biorestoration and interdiction wells
and surface treatment to achieve ground water remediation?
(7) Can we predict clean-up efficiencies using mathematical models that incorporate rate coefficients for
target pollutants that are determined in site-specific aquifer materials?
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