&EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
EPA/600/R-92/042
March 1992
Methodologies for
Evaluating In-Situ
Bioremediation of
Chlorinated Solvents
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EPA/600/R-92/042
March 1992
METHODOLOGIES FOR EVALUATING IN-SITU BIOREMEDIATION OF
CHLORINATED SOLVENTS
by
Lewis Semprini, Dunja Grbic-Galic, Perry L. McCarty, and Paul V. Roberts
Department of Civil Engineering, Stanford University
Stanford, California 94305
Cooperative Agreement EPA CR 815816
Project Officers:
Wayne C. Downs and Stephen G. Schmelling
Processes and Systems Research Division
Robert S. Kerr Environmental Research Laboratory
Ada, Oklahoma 74820
ROBERT S. KERR ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
ADA, OKLAHOMA 74820
Protection Agency
vrd, 12th Floor
. ..j'jO
Printed on Recycled Paper
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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 Cooperative Agreement EPA CR 815816 to Stanford
University. It has been subjected to the Agency's peer and administrative review, and it has been
approved for publication as an EPA document. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.
QUALITY ASSURANCE STATEMENT
All research projects making conclusions or recommendations based on environmentally related
measurements and funded by the Environmental Protection Agency are required to participate in the
Agency Quality Assurance Program. This project summarizes work that was conducted under an
approved Quality Assurance Project Plan. The procedures specified in this plan were used.
Information on the plan and documentation of the quality assurance activities and results are
available from the Principal Investigator.
11
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FOREWORD
EPA is charged by Congress to protect the Nation's land, air and water systems. Under a mandate
of national environmental laws focused on air and water quality, solid waste management and the
control of toxic substances, pesticides, noise and radiation, the Agency strives to formulate and
implement actions which lead to a compatible balance between human activities and the ability of
natural systems to support and nurture life.
The Robert S. Kerr Environmental Research Laboratory is the Agency's center of expertise for
investigation of the soil and subsurface environment. Personnel at the Laboratory are responsible
for management of research programs to: a) determine the fate, transport and transformation rates
of pollutants in the soil, the un saturated and the saturated zones of the subsurface environment;
b) define the processes to be used in characterizing the soil and subsurface environment as a
receptor of pollutants; c) develop techniques for predicting the effect of pollutants on ground
water, soil, and indigenous organisms; and d) define and demonstrate the applicability and
limitations of using natural processes indigenous to the soil and subsurface environment, for the
protection of this resource.
This report summarizes the results of several research projects where methodologies were devel-
oped for evaluating in-situ bioremediation of chlorinated solvents, which are widely encountered as
groundwater pollutants.
Clinton W. Hall
Director
Robert S. Kerr Environmental
Research Laboratory
111
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ABSTRACT
This report summarizes the behavior of and requisite conditions for a class of natural biological
processes that can transform chlorinated aliphatic compounds. These compounds are among the
most prevalent hazardous chemical contaminants found in municipal and industrial wastewaters,
landfills and landfill leachates, industrial disposal sites, and groundwater. Biological degradation
is one approach that has the potential for destroying hazardous chemicals so that they can be
rendered harmless for all time. Methodologies are presented that are useful for evaluating the
potential for biorestoration of groundwater contaminated with chlorinated aliphatic compounds.
The report is composed of six sections. Section 1 provides an introduction and an overview of the
problems with chlorinated aliphatic compounds in groundwater. Section 2 presents a review of the
processes affecting the movement and fate of chlorinated aliphatics in the subsurface, including
advection, dispersion, sorption and relative mobility, diffusional transport, and immiscible
transport. Section 3 provides a thorough review of the microbial transformation of organic
pollutants. Basic microbial metabolic processes are reviewed, focusing on anaerobic and aerobic
transformations of chlorinated aliphatic compounds. Laboratory studies of aerobic cometabolic
transformation and degradation of TCE by methanotrophs and methanotrophic communities are
summarized. In Section 4 transport and microbial process models are presented and incorporated
into a model for the aerobic cometabolic transformation of chlorinated aliphatics by methanotrophic
communities. Section 5 presents pilot-scale results of enhanced in-situ biotransformation of
halogenated alkenes, including TCE, cis- and trans-DCE, and vinyl chloride by methanotrophic
bacteria along with model simulations of the results. Section 6 presents an example study to
evaluate the potential and limitations for groundwater bioremediation at a Superfund site by
methanotrophs. Methodologies and results are presented for evaluating the presence of a native
methanotrophic community and its ability to degrade the contaminants of concern; determining the
sorption of contaminants to the aquifer material; and preliminary designing of an in-situ treatment
approach using the model previously described.
IV
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CONTENTS
Foreword iii
Abstract iv
Figures vi
Tables k
Acknowledgments x
1. Introduction 1
Purpose and overview 1
Chlorinated aliphatic compounds 2
2. Processes Affecting Movement and Fate 5
Overview 5
Advection 5
Dispersion 6
Sorption and relative mobility 8
Diffusional rate limitations 12
Immiscible transport 13
3. Biotransformation 15
Microbial transformation of organic pollutants 15
Anaerobic transformations 19
Aerobic microbial transformation of Ci and C2 chlorinated aliphatic
hydrocarbons 22
Aerobic transformation and degradation of TCE 24
4. Process Models 34
Introduction 34
Microbial processes 36
Coupling with transport processes 38
5. Results of a Pilot-Scale Study of Enhanced Biotransformation of Halogenated
Alkenes by Methanotrophic Bacteria 42
Moffett Field study 42
Model interpretation 52
6. Feasibility Studies for a Site ..51
Introduction and objectives 57
Procedures 58
Column microcosm results 62
Remediation scenarios 65
References 73
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FIGURES
Number Page
1 Variation of dispersivity with distance 7
2 Sequential breakthrough of solutes at an observation well during the Palo Alto
groundwater recharge study 9
3 Breakthrough responses for chloride, carbon tetrachloride, and tetrachloro-
ethylene at time-series sampling points a, b, c during the Borden transport
experiment 9
4 Chromatographic separation of spatial plumes at Borden 10
5 Correlation between the distribution coefficient and the octanol water partition
coefficient 12
6 DNAPL infiltration schematic 14
7 Chemical and biological transformation pathways of selected chlorinated aliphatic
compounds under anaerobic conditions 21
8 Relative rates of oxidation and reduction of a range of Ci and C2 chlorinated
compounds 23
9 The influence of formate addition on TCE transformation rates in the absence of
methane, in the pure culture Methylomonas sp. MM2 (A) and mixed culture
MM1 (B), both derived from the Moffett Field groundwater aquifer 27
10 Production (from TCE) and subsequent oxidation of CO by Methylomonas sp.
MM2 [(B) is an expanded view of the CO data from (A)] 29
11 Competitive inhibition of TCE oxidation in Methylomonas sp. MM2 by CO 30
12 The effect of TCE oxidation on subsequent methane utilization by Methylomonas
sp. MM2 32
13 Conceptual diagram of above-ground well casing and aquifer (in-situ) bioreactor
35
14 Conceptual model of transport and biodegradation processes that must be
considered 35
VI
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Number Page
15 Michaelis-Menten enzyme kinetics 36
16 Mass transfer considerations 40
17 A vertical section of the test zone 43
18 Schematic of the automated Data Acquisition and Control system 43
19 Map of the well field installed at the field site 44
20 Bromide tracer breakthrough and elution in the TracerS experiment 45
21 Schematic of the chemical injection system 48
22 Methane and DO response at the S2 observation well due to the biostimulation of
the test zone 48
23 Decreases in normalized concentration of vinyl chloride, trans-DCE, and cis-
DCE at the S2 well in response to biostimulation in the third season 49
24 Decreases in normalized concentration of vinyl chloride, trans-DCE, and cis-
DCE at the S1 well in response to biostimulation in the third season 51
25 Response of trans-DCE and cis-DCE at the S1 well to the injection of
(1) methane, (2) formate, (3) methane and formate, (4) methanol,
and (5) no electron donor 51
26 Model simulation and observed methane and DO response at the S2 observation
well during the first season of field testing 54
27 Predicted biomass concentration at a node 2.2 m from the injection well due to
stimulation with short and long pulses 54
28 Simulations of the response of methane, VC, t-DCE, and c-DCE, at the S2 well
to biostimulation of the test zone in the third season of field testing 55
29 Model simulations and the response of methane, VC, t-DCE, and c-DCE at the
SI well using competitive inhibition kinetics and rate-limited sorption 55
30 Distribution of TCE, DCE, and VC at the St. Joseph site 58
31 Vertical profile of subsurface at St. Joseph 59
32 Schematic diagram of column microcosms 60
33 Schematic figure of procedure used in exchanging column fluids 61
34 Breakthrough curves of bromide tracer for the different columns 61
35 Methane removal in column microcosms 62
vii
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Number Page
36 VC, t-DCE, and TCE removal in column microcosms 64
37 In-situ bioremediation case simulated 67
38 Pump-and-treat system for comparison with bioremediation 67
39 Simulation response to biostimulation with methane for VC remediation 69
40 Comparison of in-situ bioremediation of VC with pump-and-treat.
Biostim+pump is a combination of surface treatment and in-situ treatment 69
41 Comparison of in-situ bioremediation of t-DCE with pump-and-treat 71
42 Comparison of in-situ bioremediation of c-DCE with pump-and-treat 71
Vlll
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TABLES
Number Page
1 Frequency of Occurrence of Volatile Organic Chemicals 2
2 Common Halogenated Aliphatic Contaminants 4
3 Distribution Coefficients and Retardation Factors 11
4 Most Prevalent Chemical Compounds at U.S. Superfund Sites with a Specific
Gravity Greater than Unity 14
5 Classification of Microorganisms by Major Catabolic Requirements 16
6 Classification of Microbial Physiological Groups by Catabolic Electron Acceptor 17
7 Types of Organic Pollutant Transformation According to the Role of the
Pollutant in Microbial Metabolism 18
8 Transformations of Halogenated Aliphatic Compounds 20
9 Kinetic Parameters for TCE Transformation by the Pure Culture Methylomonas
sp. MM2 Grown in the Presence or Absence of a Metal Chelator, EDTA 31
10 Summary of the Important Factors for in-Situ Treatment of Chlorinted Solvents
by Methanotrophic Communities 33
11 Selected Shallow Biofilm Models for Microbial Transformation in Porous Media 39
12 Selected Biofilm Models for Microbial Transformations 39
13 Comparison of Bromide Tracer Tests Under Induced Gradient Conditions 46
14 Residence Times and Retardation Factors for the Chlorinated Organic Compounds
Based on the Time Required to Achieve 50% Fractional Breakthrough 46
15 Extent of Biotransformation-Third Field Season 50
16 Model Parameters for Simulation of Chlorinated Organics in Biostim3 Shown in
Figure 29 56
17 Engineering Summary of in-Situ Biological Treatment 72
IX
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ACKNOWLEDGMENTS
The information presented here integrates research work performed over the past 5 years in two
projects sponsored by the Robert S. Kerr Environmental Research Laboratory of the U.S.
Environmental Protection Agency, CR-812220 and CR-815816. The Feasibility Study at the St.
Joseph field site was supported with financial assistance from Allied-Signal, Inc., Bendix
Automotive System-North America, and the U.S. EPA-sponsored Western Region Hazardous
Substance Research Center, Grant No. CR-815738.
Individuals who have contributed to these studies include: Lisa Alvarez-Cohen, Constantinos
Chrysikopoulos, Mark Dolan, Steve Gorelick, Franziska Haag, Thomas Harmon, Susan Henry,
Gary Hopkins, Robert Johns, Nancy Lanzarone, Douglas Mackay, Kevin Mayer, Martin
Reinhard, Robert Roat, Claire Tiedeman, and Timothy Vogel.
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SECTION 1
INTRODUCTION
PURPOSE AND OVERVIEW
Chlorinated aliphatic compounds are among the most prevalent hazardous chemical contaminants
found in municipal and industrial wastewaters, landfills and landfill leachates, industrial sludges,
waste disposal sites, and groundwater. Several of these chemicals have been widely used as sol-
vents for household, commercial, and industrial cleaning and degreasing operations. Included here
are carbon tetrachloride (CT), tetrachloroethylene (PCE), trichloroethylene (TCE), 1,1,1-
trichloroethane (TCA), and methylene chloride (MC). In the past, used solvents were often dis-
charged indiscriminately into the environment. A characteristic of these man-made solvents is their
relative resistance to breakdown by natural microorganisms, and hence they have tended to persist
and accumulate in the environment. While efforts are now being made to clean with other less
harmful and less environmentally persistent chemicals, or to clean and recycle used chlorinated
solvents, there is a legacy of chlorinated-solvent-contaminated environments that need to be reme-
diated.
Most current cleanup methods for chlorinated aliphatic compounds employ physical processes
which tend to transfer the compounds to another medium or to concentrate the compounds for
burial elsewhere. For example, groundwater contaminated by chlorinated solvents is often
pumped to the ground surface and air stripped, which sends the solvents into the atmosphere.
Now that this method is being restricted in some areas, the air stripped solvents sometimes are
sorbed and concentrated onto activated carbon, which then may be hauled to a hazardous chemical
disposal site. In such schemes the chemicals are simply moved from one part of the environment
to another, where they may cause other problems. Biological decomposition is one approach that
has the potential for destroying hazardous chemicals so that they are rendered harmless for all time.
This natural process is widely used for more easily biodegraded compounds, especially where
chemicals exist in relatively dilute form (mg/kg levels) in waste water or sludges, or in the envi-
ronment. Although chlorinated aliphatic compounds have been known for quite some time to be
resistant to normal biological transformations, many natural organisms have now been found that
can bring about their decomposition. This raises the possibility that these chemicals can also be
destroyed by engineered biological processes.
This report provides an overview of the characteristics of the natural biological processes that can
transform chlorinated aliphatic compounds. It also describes research and field studies that have
been directed towards taking advantage of these natural processes for in-situ restoration of
groundwaters. Methodologies are presented that might be applied for evaluating the potential for
biorestoration of contaminated groundwater. An example study to evaluate the potential and limi-
tations for groundwater bioremediation at a Superfund site is also included.
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CHLORINATED ALIPHATIC COMPOUNDS
Table 1 contains a listing of the reported frequency of occurrence of several volatile organic chemi-
cals (VOCs), which includes the halogenated aliphatic compounds of interest here. The four most
prevalent VOCs in groundwater used as a source of drinking water supply are the trihalomethanes,
which are formed from chlorination for disinfection. As such, these compounds are not major
contaminants of the groundwater itself. The trihalomethanes are also among the dominant VOCs in
treated municipal wastewater, again a result of chlorination of drinking water supplies. The next
most prevalent VOCs are the three major chlorinated solvents PCE, TCE, and TCA. Not far down
on the list is the chlorinated solvent CT, which has seen decreased usage in recent years but was a
major solvent in the past.
TABLE 1. FREQUENCY OF OCCURRENCE OF VOLATILE ORGANIC CHEMICALS
Order of Frequency
voc
Chloroform
Bromodichloromethane
Dibromochloromethane
Bromoform
Tetrachloroethylene
Trichloroethylene
1,1,1 -Trichloroethane
1 , 1 -Dichloroethane
1 ,2-Dichloroethylene
Carbon tetrachloride
1 , 1 -Dichloroethylene
m-Xylene
o,p-Xylene
Toluene
p-Dichlorobenzene
1 ,2-Dichloropropane
Dichloroiodomethane
Benzene
Ethylbenzene
Bromobenzene
Vinyl chloride
Groundwater0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
Treated
Municipal
Wastewatei*
3
6
7
9
2
5
1
8
11
4
10
CERCLA
Sites'
7
4
1
6
9
12
15
10
5
15
3
11
13
2
7
8
a Based upon survey of treated groundwater used for drinking-water Supply (Westrick et al.,
1984).
b Survey of four treatment plants in Arizona, California, and Colorado (McCarty, 1990).
c U.S. EPA, 1990.
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In contaminated groundwater, another group of chlorinated aliphatic compounds appear that are not
generally prevalent in municipal wastewaters. These are 1,1-dichloroethane (1,1-DCA), 1,2-
dichloroethylene (1,2-DCE), 1,1-dichloroethylene (1,1-DCE), and somewhat further down the
list, vinyl chloride (VC). These compounds often are produced slowly in the environment under
appropriate conditions as daughter products from chemical and biological transformations of chlor-
inated solvents, which are the original contaminants.
Another group of prevalent VOCs consist of the aromatic hydrocarbons: benzene, xylene, and
ethylbenzene. These compounds represent the soluble components of gasoline and are indicators
of contamination by petroleum products. These compounds, however, are not the subject of this
report.
Table 2 provides a listing of the major one- and two-carbon halogenated VOCs that are considered
in this report, the chemical formulas for each, acronyms that are used, and U.S. drinking water
maximum contaminant limits (MCL) that have been established. Federal MCLs are not available
for compounds without MCLs specified in Table 2, but some states have established regulations
for at least some of them. Reaching the low contaminant levels represented by the federal drinking
water standards presents a significant challenge for any bioremediation process, especially when
contaminants occur two to four orders of magnitude higher in concentration at many contaminated
sites of concern.
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TABLE 2. COMMON HALOGENATED ALIPHATIC CONTAMINANTS
U.S. Drinking-
WaterMCL
Chemical
Formula
Acronym
Trihalomethanes
Chloroform
Bromodichloromethane
Dibromochloromethane
Bromoform
Other Chlorinated Methanes
Carbon tetrachloride
Methylene chloride
Chlorinated Ethenes
Tetrachloroethylene
Trichloroethylene
cis- 1 ,2-Dichloroethylene
trans- 1 ,2-Dichloroethylene
1 , 1 -Dichloroethylene
Vinyl chloride
Chlorinated Ethanes
1,1,1 -Trichloroethane
1 , 1 -Dichloroethane
1 ,2-Dichloroethane
Chloroethane
CHC13
CHBrCl2
CHBr2Cl
CHBr3
ecu
CH2C12
cic _ cci
all
f*1 — f^
HTT
c* — c*
C1C = CH
Cl/-i f~H
a\^ — VxTT
tl
all
p O
CC13CH3
CHC12CH3
CH2C1CH2C1
CH2C1CH3
CF
)
CT
MC
PCE
TCE
C-1.2-DCE
t-l,2-DCE
1,1 -DCE
VC
TCA
1,1 -DCA
1,2-DCA
CA
S = 100
5
5
5
70
100
7
2
200
5
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SECTION 2
PROCESSES AFFECTING MOVEMENT AND FATE
OVERVIEW
Once released onto the land surface or underground, synthetic organic chemicals, such as the halo-
genated aliphatics, are subject to a variety of influences that lead to an extraordinarily complex
pattern of behavior. The processes influencing the transport, distribution and fate of these chemi-
cals include the following:
1) Advection, the miscible transport in aqueous solution under the influence of the hydraulic
potential gradient;
2) Dispersion, the mixing and spreading of concentration fronts, that arises largely from differ-
ential rates of movement along the myriad individual flow paths through the porous medium;
3) Sorption, the partitioning of a compound between the moving solution and the stationary
solid phase;
4) Immiscible transport, the migration of slightly soluble chemicals as a separate liquid phase,
often driven downward by density difference in the case of halogenated aliphatics; and
5) Diffusional transport, the slow migration of solute molecules into the matrix rock or dead-end
pores under the influence of a concentration driving force.
The influence of these factors on contaminant behavior has been summarized in several recent
reviews (McCarty et al., 1981; NAS, 1984; Anderson, 1984; Mackay et al., 1985; Goltz and
Roberts, 1986). The present section focuses on the principles underlying the transport, distribu-
tion, and fate of halogenated aliphatic compounds; transformation processes are described in
Section 3. Field studies of organic contaminant behavior are emphasized, as these often reveal the
surprising significance of phenomena that occur in real subsurface environments over long time
periods in a way that could not be demonstrated from theoretical and laboratory studies alone
(Roberts et al., 1982b, 1986; Anderson, 1987).
ADVECTION
The transport of nonreactive, dissolved contaminants is driven by the same hydrostatic forces
responsible for groundwater movement. The governing principles have been expounded in stan-
dard references on groundwater movement, such as Freeze and Cherry (1979), where methods of
investigation are also summarized. In most instances of groundwater flow through fine- to
medium-grained porous media, Darcy's Law applies:
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v - ^l^
v = ~KdL = vn (1)
where v = specific discharge or Darcy velocity, an average velocity based on the total cross-
sectional area [L/t], v = average linear velocity [L/t], n = porosity [-], K = hydraulic conductivity
[LA]; and dh/dL = hydrauh'c gradient [-].
The porosity, n, represents the intergranular pore volume expressed as a fraction of the total void
volume. The value of n depends on the grain shape and size distribution, with typical values on
the order of n = 0.25 to 0.5 for silt, sand, and gravel, and somewhat higher values, n = 0.4 to 0.7
for clay.
The hydraulic conductivity, K, depends mainly on the grain size of the porous medium, although
fluid density and viscosity play a minor role, according to
where C = empirical drag coefficient [-], d = grain diameter [L], p = fluid density [M/L3], (I = vis-
cosity, and k = intrinsic permeability = Cd2. Accordingly, both the intrinsic permeability and the
hydraulic conductivity are strong functions of the grain diameter, increasing approximately as d2.
Approximate values of the hydrauh'c conductivity and intrinsic permeability can be found in Table
2.2 of Freeze and Cherry (1979). For example, the hydrauh'c conductivity for a clean sand usually
lies between 10~4 and 1 crn/s, whereas values for silty sand range about an order of magnitude
lower. The magnitude of the hydrauh'c conductivity is significantly reduced by the presence of fine
material, which tends to fill the interstices between the coarser grains, thus reducing the effective
pore size.
Under natural conditions, the hydraulic gradient is influenced by the presence of recharge sources
and discharge sinks as well as their separation distances and relative elevations. In areas of rela-
tively flat topography, values of the hydraulic gradient often are in the range from 10"3 to 10'5,
with higher values encountered in areas of high recharge and steep slope. The hydraulic conduc-
tivity also exerts an indirect effect on the hydraulic gradient, introducing a complexity that some-
times greatly complicates data interpretation and modeling. However, the hydraulic gradient is
influenced significantly by injection or extraction wells, which may increase the hydraulic conduc-
tivity by several orders of magnitude.
Considering the combined effects of hydraulic conductivity, hydraulic gradient, and porosity, it
can be expected that linear velocities of groundwater transport in uniform sand and gravel aquifers
in areas of gentle topography are on the order of 1 to 1000 m/yr (Mackay et al.,1985). Values in
this range, frequently on the order of 10 to 100 m/yr, are often observed at hazardous waste sites
in such settings. Thus, groundwater movement is relatively slow in most instances, a fact of
salient importance in remediation efforts.
DISPERSION
Dispersion refers to the sum effect of processes that tend to spread an advancing solute concentra-
tion wave or pulse. The underlying processes are thought to be the following: molecular diffu-
sion, hydromechanical mixing, and differences in hydraulic conductivity along different flow
paths. The dispersive mixing can be longitudinal, i.e., in the direction of flow, or transverse, i.e.,
-------
orthogonal to the flow. Thus, dispersion tends to dilute contaminant concentrations by spreading
the contaminant, but also hastens the arrival of concentration fronts.
The extent of dispersion is highly dependent on the heterogeneity of the porous medium, because a
wide range of hydraulic conductivities results in large spatial variation of velocities along different
flow paths. The dispersion coefficient, D [L2/t], the basic rate coefficient governing the extent of
dispersive mixing, can be estimated as the product of the fluid velocity, v [L/t], and a dispersive
mixing length scale, a [L]
D = v • a (3)
The dispersivity, a, captures the effects of the porous medium's heterogeneity, and must be esti-
mated by interpretation of tracer experiments. Estimates of a based on field measurements typi-
cally exceed those from laboratory data by several orders of magnitude, owing to the greater
heterogeneity of the field settings (Freeze and Cherry, 1979). Moreover, comparison of estimates
based on field measurements at different scales shows that the dispersivity estimates usually
increase with increasing scale of observation. Anderson (1984) suggests that the estimated dis-
persivity increases approximately in proportion to the scale of observation, as shown in Figure 1.
100
CO
QL
LU
Q_
£ 'o
t I
SAND, GRAVEL,
* SANDSTONE
A LIMESTONE, BASALT,
GRANITE ft SCHIST
10 100 1000
DISTANCE (m)
Figure 1. Variation of dispersivity with distance (from Anderson, 1984).
7
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The values of a for a given observation distance, Z, are distributed over an order of magnitude
above and below the fitted straight line in Figure 1, presumably because some sites exhibit consid-
erably more heterogeneity than others.
The extent of dispersive mixing or spreading can be quantified in terms of a dimensionless group
known as the Peclet Number, Pe = X/a, where X is the integral length scale of observation in the
longitudinal direction. The larger the Peclet Number, the smaller the dispersive length scale com-
pared to the total travel distance, and hence the smaller the role of dispersion relative to advection.
If Pe > 100, dispersion exerts a minor influence on contaminant transport and distribution,
whereas if Pe < 10, dispersion has a major effect. If the linear relationship between log a and log
X in Figure 1 is valid, then the values of Pe for the collection of field sites fall within an order of .
magnitude of Pe = 10, suggesting that dispersion is of moderate importance at typical field sites.
However, it can be fairly stated that dispersion does not directly influence the fate of contaminants
in solution, since it acts only to spread out the contaminant mass over a greater volume.
SORPTION AND RELATIVE MOBILITY
The halogenated aliphatic compounds do not sorb to soils and aquifer materials as readily as do
many pesticides, but nevertheless, sorption in aquifer systems is sufficient to retard the rate at
which they move in groundwater in relation to the movement of groundwater itself. This relative
movement can be expressed mathematically by the retardation equation (Freeze and Cherry, 1979):
vc
where v = average linear velocity of groundwater, vc = average linear velocity of the contaminant,
Pb = mass density of solids in aquifer, n = porosity, and Kd = distribution coefficient.
The term (1 + pblQj/n) is commonly known as the retardation factor. For aquifer materials, pb is
approximately 1.7 g/cm^, and n generally varies between 0.2 and 0.4 (Freeze and Cherry, 1979).
With these units, K^ should be expressed in units of cm3/g. K
-------
to
*-
1
05
*
u.
- CHLORIDE
CHLOROFORM
BROMOFORM
CHLOROBENZENE
10 20
VOLUME hO'm3)
30
Figure 2. Sequential breakthrough of solutes at an observation well during the Palo Alto
groundwater recharge study (Roberts et al., 1982b).
i
0.9
o.a
0.7
o.a
t
°-9
I o.*
' 0.3
0.2
0.1
0
— CMoride
+ Caftan tetracrtoride
o Tetradtoro0thytan0
200
me (DAYS)
•»oo
Figure 3. Breakthrough responses for chloride, carbon tetrachloride, and tetrachloroethylene at
time-series sampling points a, b, c during the Borden transport experiment (Roberts et
al., 1986).
-------
X
(m)
60
50
40
30
20
10
-10
CTET
633 days
PCE
633 days
-10
10
20
30
y(m)
Figure 4. Chromatographic separation of spatial plumes at Borden: chloride (Cl, 647 days);
carbon tetrachloride (CTET, 633 days); and tetrachlorethylene (PCE, 633 days).
Contour interval depicted for Cl is 5 mg/1 beginning at an outer contour of 10 mg/1.
Contour intervals depicted for CTET and PCE are 0.1 |lg/l beginning at an outer
contour of 0.1 p.g/1 (Roberts et al., 1986).
1987). The aquifer material at Borden had a low organic content (0.02%), and consisted of quartz
(58%), amphiboles (7%), feldspars (19%), carbonates (14%), and miscellaneous materials (2%).
Measured values of K
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the distribution coefficients measured in the laboratory (Curtis et al., 1986a), it was found that the
retardation factor appeared to increase with time and distance from the point of injection. This
appears to be related to a slow rate of diffusion of the contaminants into the aquifer solids, the
characteristic time scale for which can be measured in terms of weeks to months rather than hours
as commonly assumed (Ball and Roberts, 1991b). To measure sorption equilibrium with such
extended contact periods without introducing artifacts, a special protocol employing flame-sealed
ampules has been developed (Ball and Roberts, 1991a); alternatively, aquifer materials can be
pulverized to expedite the equilibrium measurements.
TABLE 3. DISTRIBUTION COEFFICIENTS AND RETARDATION FACTORS
Retardation
CTET
BROM
PCE
DCB
HCE
logioKow
2.7
2.3
2.6
3.4
3.6
IQxlO6
(m3/g)
0.15
0.17
0.45
0.76
0.81
Predicted0
1.3
1.2
1.3
2.3
2.3
Batch
Experiments^
1.9
2.0
3.6
6.9
5.4
Temporal
1.6-1.8
1.5-1.8
2.7-3.9
1.8-3.7
4.0
Spatial
1.8-2.5
1.9-2.8
2.7-5.9
3.9-9.0
5.1-7.9
a Calculated from the regression by Schwarzenbach and Westall (1981) with foc = 0.02%.
b Curtis et al. (1986a).
c Roberts et al. (1986).
While others have shown a relatively good correlation between K^ and the aquifer organic content
and the contaminant's octanol/water partition coefficient, KOW (Karickhoff et al., 1979;
Schwarzenbach and Westall, 1981), this correlation is relatively poor for aquifers with low organic
content (foe < 0.1%, McCarty et al., 1981), and for similar chemicals as represented by the chlori-
nated aliphatic compounds summarized. Generally aquifers are fairly poor in organic matter con-
tent so that the retardation noted appears to be more a function of sorption to inorganic rather than
organic materials (Curtis et al., 1986b). Figure 5 (Curtis et al., 1986a) indicates that the correla-
tion between the distribution coefficient normalized to the organic content (Koc = K^/foc) and the
octanol/water partition coefficient is relatively poor within this group of compounds: bromoform
(BROM), tetrachloroethene (PCE), carbon tetrachloride (CTET), 2-dichlorobenzene (DCB), and
hexachloroethane (HCE), in order of increasing KQW.
Retardation is an important process in groundwaters for at least two reasons. Firstly, since chemi-
cals have different retardation coefficients, their relative rates of movement through aquifers will
differ widely (Roberts et al., 1982a). Thus, if an aquifer were to become contaminated with
several compounds at one location, they would tend to move at different speeds and would arrive
at a downgradient well at different times. This would make the contaminant composition of the
extracted water different than that at the point of contamination, often making it difficult to verify
the original source of contamination. The other important aspect of retardation is that, as noted,
retardation provides a basis for estimating the relative amount of the contaminant present in the
aqueous phase as compared with that sorbed to the aquifer solids. For example, for a retardation
factor of 5, one-fifth of the contaminant would be present in the aqueous phase and four-fifths
would be sorbed onto the aquifer solids. Restoration of a contaminated aquifer would require that
11
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4.5
4.0
O
O
o>
3.0
2.5
2.
Log Hoc • 0.51 Log Ko» • 1.79
Log Koe • 0. 72 Log How • a 49
(Sch»ori«nbach and Vaatall 1981)
£
2. 5
3. 0 3. 5
Log Kow
4. 0 4. 5
Figure 5. Correlation between the distribution coefficient and the octanol water partition
coefficient (from Curtis et al., 1986a).
the contaminants be removed from the solid phase as well as from the aqueous phase. Also,
sorption is an important factor in damping concentration fluctuations that arise as packets of water
with differing composition move through an aquifer; sorption tends to strengthen the tendency of
such fluctuations to be smoothed during groundwater transport (Roberts and Valocchi, 1981;
Valocchi and Roberts, 1983). Finally, sorption of contaminant may also affect contaminant trans-
formation either by hindering or accelerating the transformation process. For these various
reasons, sorption is an important process in aquifers; and more basic information is needed about
factors affecting it.
DIFFUSIONAL RATE LIMITATIONS
Mass transport between the moving groundwater and the stationary solids is sufficiently rapid
under ideal circumstances to assure equilibrium distribution of contaminants between the pore
water and solids. However, both theory and recent field evidence suggest strongly that the fate of
organic contaminants in natural hydrogeologic settings may depend critically on rates of transport
between the pore water and stationary solid phases. For example, a field-scale study suggested
that organic contaminant sorption equilibrated very slowly, over a period of several years (Roberts
et al., 1986; Goltz and Roberts, 1986). Diffusional rate limitations are especially likely to arise
under circumstances where the sorption capacity of the solids is large and the porous medium is
heterogeneous, containing extensive zones of low permeability (Brusseau et al., 1989; Valocchi,
1985). Serious rate limitations can occur at the grain scale even in relatively homogeneous media,
if the sorbing solids are microporous (Ball and Roberts, 1991a,b) or aggregated (Rao et al., 1982).
The implications for aquifer remediation are obvious, and mostly adverse (Criddle et al., 1991;
Reijnaarts et al., 1990). Deviations from local equilibrium can lead to situations in which 1) a
12
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large proportion of sorbed contaminant remains in the solid phase long after aqueous concentra-
tions have subsided to low levels as a result of purging with clean fluid, and 2) the local solution
concentrations are governed by the slow release of previously sorbed contaminants. These compli-
cations are likely to hinder both pump-and-treat and in situ remediation efforts, by substantially
reducing the rate of contaminant removal by either purging or reaction, thus extending the time
duration of remediation efforts. Such conditions are especially likely to exert a significant impact
on remediation performance if 1) the contaminants sorb strongly, and/or 2) the porous medium is
significantly heterogeneous.
It would be premature, given present knowledge, to propose quantitative criteria for deciding
whether diffusional rate limitations will affect remediation significantly. Judging from field
observations of contaminant behavior at the Borden and Moffett Field sites, we surmise that
retardation factors greater than approximately three can lead to significant rate limitations if a
substantial portion of the sorption capacity is associated with zones having permeability signifi-
cantly (e.g., more than an order of magnitude) below average (Roberts et al., 1986; Semprini and
McCarty, 1992). These conditions are believed to apply to many, if not most, remediation situa-
tions. Hence there is an urgent need to improve methods for characterizing remediation sites with
respect to the factors controlling the rates of sorbed contaminant release from solid phases, and to
employ those methods systematically in site investigations.
IMMISCIBLE TRANSPORT
Transport of immiscible liquid phases, known colloquially as non-aqueous phase liquids or
NAPLs, has come to be recognized as a major route of contaminant spreading at hazardous waste
sites and as a significant factor affecting contaminant distribution (Ruling and Weaver, 1991). The
principles of immiscible contaminant movement and the role of contaminant properties have been
reviewed recently (Mercer and Cohen, 1990; Ruling and Weaver, 1991). The halogenated
aliphatic compounds emphasized in this document are susceptible to pronounced vertical transport
as immiscible phases: almost all are specifically dense, slightly miscible liquids under conditions
typical of the subsurface, with solubilities on the order of 0.1 to 10 g/1 and specific gravities on the
order of 1.1 to 2 [Table 4 (from Ruling and Weaver, 1991)]. Such compounds that are more
dense than water are commonly referred to as DNAPLs. As shown by Schwille (1988) in physical
model experiments, compounds of this kind tend to migrate downward through the vadose and
groundwater zone, driven by gravity forces which dominate the hydrostatic forces that impel
groundwater movement and solute transport (Figure 6).
In contrast to groundwater transport, DNAPL movement is primarily vertical, except where the
descending immiscible organic liquid encounters a barrier posed by a stratum of low permeability.
Hence, the DNAPL transport is uncoupled from the usual mode of advective solute transport in
several ways: with respect to direction, velocity, and pathways. Research on the transport and
distribution of DNAPLs in porous media is still in its infancy. Most investigations have centered
on relatively homogeneous glass bead and sand packs; and it is altogether uncertain how immis-
cible liquids behave in markedly heterogeneous media, especially where layers or fractures
abound. Further research is needed to develop practical means of characterizing hazardous waste
sites from the viewpoint of processes governing immiscible liquid behavior before the ramifica-
tions for bioremediation can be assessed realistically.
13
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TABLE 4. MOST PREVALENT CHEMICAL COMPOUNDS AT U.S.
SUPERFUND SITES WITH A SPECIFIC GRAVITY GREATER THAN UNITY
(from Huling and Weaver, 1991)
Water
Solubility
Compound
Halogenated Volatiles
Density
(mgfl)
Chlorobenzene
1 ,2-Dichloropropane
1,1-Dichloroethane
1 ,2-Dichloroethylene
1 ,2-Dichloroethane
Trans-1 ,2-Dichloroethylene
Cis- 1 ,2-Dichloroethylene
1,1,1 -Trichloroethane
Methylene Chloride
1 , 1 ,2-Trichloroethane
Trichloroethylene
Chloroform
Carbon Tetrachloride
1 , 1 ,2,2-Tetrachloroethane
Tetrachloroethylene
Ethylene Dibromide
1.1060
1.1580
1.1750
1.2140
1.2530
1.2570
1.2480
1.3250
1.3250
1.4436
1.4620
1.4850
1.5947
1.6
1.6250
2.1720
4.9
2.7
5.5
4.0
8.7
6.3
3.5
9.5
1.3
4.5
1.0
8.2
8.0
2.9
1.5
3.4
E+02
E+03
E+03
E+02
E+03
E+03
E+03
E+02
E+04
E+03
E+03
E+03
E+02
E+03
E+02
E+03
ONAPL
AIR OR WATER -
FILLED PORE SPACE
TOP OF
CAPILLARY FRINGE
WATER TABLE
WATER-FILLED
PORE SPACE
Figure 6. DNAPL infiltration schematic (from Mercer and Cohen, 1990).
14
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SECTION 3
BIOTRANSFORMATION
MICROBIAL TRANSFORMATION OF ORGANIC POLLUTANTS
Microbiological processes have a great significance in determining the fate and transport of envi-
ronmental pollutants. Chemoorganotrophic bacteria and fungi are especially prominent in this
respect (Alexander, 1980; Bumpus et al., 1985). These microorganisms produce natural catalysts
(enzymes) which break down organic substrates to obtain carbon, energy, reductants, and some-
times also nutrients necessary for growth. Frequently, these degradative (catabolic) enzymes can
also transform (i.e., partially change), or degrade (i.e., completely destroy) organic pollutants.
The degree of chemical change of the pollutant molecule may vary and will depend on the chemical
and physical properties of this compound, as well as on the physiological types of microorganisms
involved, and physical and chemical environmental factors (Boethling and Alexander, 1979;
Alexander, 1981, 1985).
The outcome of the pollutant-microorganism interaction is affected by the environment to such an
extent that no transformation will occur if the environment is not conducive to microbial growth, or
if it renders the pollutant unavailable to microorganisms. The most important environmental vari-
ables are the following: presence/absence of microbial growth substrates, electron acceptors,
nutrients, inhibitors, other microorganisms, and toxic compounds; water activity; ionic strength;
chemical composition of the medium; pH; oxidation-reduction potential; temperature; and adsorp-
tive surfaces (Alexander, 1965; Morrill et al., 1982).
Some pollutants, however, especially xenobiotics (man-made chemicals with novel structures,
Leisinger, 1983), persist for long periods of time in various types of environments, or are intrinsi-
cally nonbiodegradable because of their physical and chemical properties. Such compounds are
termed recalcitrant pollutants (Alexander, 1973). The notion of recalcitrance and the classification
of substances in that regard are subject to change: new microorganisms are constantly being dis-
covered that exhibit degradative or transforming capabilities towards pollutants previously believed
to be recalcitrant (Brown et al., 1987), or under environmental conditions previously considered
unconducive to such transformations (Wilson and Wilson, 1985; Grbi6-Galic, 1990; Lovley and
Lonergan, 1990).
Microbial Metabolism
Provided favorable environmental conditions, microorganisms will grow and multiply, and their
population will increase. Microbial metabolism is the basis for growth. Degradative (catabolic')
reactions yield energy, reductant supply, and simple building blocks necessary for synthesis of
new cells; biosynthetic (anabolic) reactions put all of these together to synthesize cellular con-
stituents (Gottschalk, 1986). In order for the metabolism to function, several essential require-
ments need to be satisfied: electron donor, electron acceptor, carbon source, energy source, water,
and nutrients.
15
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Chemoorganotrophs (Table 5) are the most important class of microorganisms with regard to the
transformation and degradation of organic pollutants. These microorganisms frequently satisfy
several of their requirements (e.g., carbon, energy, and electron donor demands) from the same
organic compound. This organic compound is then termed primary substrate. Since the carbon
source for these organisms is organic, they are frequently called heterotrophs (because they build
their organic constituents from other organic compounds). On the contrary, chemolithotrophs.
bacteria with equally powerful enzymes but different preferences, obtain their energy and electrons
from inorganic chemicals such as ammonia, nitrite, hydrogen sulfide, sulfur, molecular hydrogen,
etc. Their carbon demand is satisfied from an inorganic source, carbon dioxide, and therefore they
are called autotrpphs (capable of synthesizing their organic carbon from an inorganic source).
Other physiological groups exist, but they are not so important in degradation of organic pollutants
(except maybe photoorganotrophs) and are not within the scope of this report.
TABLE 5. CLASSIFICATION OF MICROORGANISMS BY MAJOR CATABOLIC
REQUIREMENTS
Microorganisms
Energy Source
Carbon Source
Electron Donor
(Reductant)
Chemoorganotrophs
Chemolithotrophs
(Chemoautotrophs)
Photoorganotrophs
Photolithotrophs
(Photoautotrophs)
Organics
Inorganics
(e.g., NH+, N02
H2, Fe2+, H2S, S, etc.)
Light
Light
Organics
Carbon dioxide
(002)
Organics, CO2
Carbon dioxide
(002)
Organics
Inorganics
(e.g., NH+, N02
H2, Fe2+, H2S, S, etc.)
Organics
Water (H2O), or other
reduced inorganics
(e.g., H2, H2S, S)
Regardless of the physiological group, microorganisms also need an electron acceptor which,
when coupled to the electron donor, establishes an electron transfer (oxidation-reduction reaction)
that plays an important role in generation of useful energy (Table 6). Aerobic microorganisms use
molecular oxygen as an electron acceptor; anaerobic microorganisms respire other compounds,
such as nitrate or nitrous oxide (denitrifiers and nitrate reducers), ferric iron (iron reducers).
sulfate, thiosulfate or sulfur (sulfate reducers), protons (obligate proton-reducing acetogens). or
carbon dioxide (methanogens). Recently, microorganisms have been found which can utilize
selenate (Oremland et al., 1989) or chlorinated organic compounds such as chlorinated aromatic
acids (Dolfing and Tiedje, 1987) as electron acceptors and can obtain useful metabolic energy
through such reactions.
Finally, there are bacteria which utilize their primary substrate not only as a carbon and energy
source and an electron donor but also as an electron acceptor. In this process, which happens in
the absence of exogenous electron acceptors, an incomplete degradation of the primary substrate
occurs; some of the products are more reduced than the primary substrate, whereas the others are
more oxidized. The bacteria which carry out this type of catabolism are called fermenters.
16
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TABLE 6. CLASSIFICATION OF MICROBIAL PHYSIOLOGICAL GROUPS BY
CATABOLIC ELECTRON ACCEPTOR
Catabolic Process Electron Acceptor
Aerobic respiration Oxygen (02)
Anaerobic respirations.
Denitrification, nitrate reduction Nitrate (NOA nitrous oxide (NiO)
Sulfate reduction Sulfate (SO^), thiosulfate (8203")
Sulfur reduction Sulfur (S)
Ferric iron reduction Ferric iron (Fe3+)
Obligate proton reduction Protons (H+)
Methanogenesis Carbon dioxide (CO2)
Organic respiration Organics (e.g., fumarate)
Fermentation Organics as both donors and acceptors of electrons
Types of Pollutant Transformation (Table 71
Bacteria can frequently use numerous organic pollutants as their primary substrates (energy
sources, carbon sources, and/or electron donors). As the result, the microorganisms grow on the
pollutants, their populations increase, and consequently the rate of pollutant degradation also
increases if the biomass concentration is rate-limiting (Alexander, 1980). This is the best possible
outcome in the battle of environmental detoxification and reclamation, because it usually results in
complete degradation (mineralizationl of the pollutant molecule: an organic compound is converted
to innocuous inorganic products such as carbon dioxide, water, ammonia, chloride, etc. For
example, simple aromatic and aliphatic constituents of petroleum, and even some polynuclear aro-
matic hydrocarbons (PAH), represent excellent primary substrates for bacteria (see reviews by
Britton, 1984, and Gibson and Subramanian, 1984). Indigenous microorganisms have been suc-
cessfully used for in-situ bioreclamation of gasoline-contaminated aquifers following injection of
oxygen or hydrogen peroxide (Raymond et al., 1975) or nitrate (Battermann and Werner, 1984) as
electron acceptors for microorganisms.
Frequently, however, organic pollutants which could be used as primary substrates by bacteria are
present in such low (trace) concentrations that they cannot support microbial growth. The micro-
organisms can deal with this situation if there is another, sufficiently abundant, primary substrate
present in their environment; this substrate is utilized for growth, and the pollutant can be simulta-
neously degraded through a process termed secondary substrate utilization (Rittmann et al., 1980a;
McCarty et al., 1981). This type of pollutant degradation also usually results in mineralization.
Unfortunately, mineralization does not always happen; numerous pollutants are only partially
transformed by microorganisms. Fungi, typically, catalyze various detoxification reactions (partial
transformations) rather than catalyzing catabolic breakdown of xenobiotics (Gibson and
Subramanian, 1984). Even bacteria often catalyze only a partial chemical change which can result
in a loss of toxic properties (detoxificationX or no loss of toxicity, or even generation or amplifica-
tion of toxicity (activation). A well-known example of activation is reductive dechlorination of
tetrachloroethylene (PCE) and trichloroethylene (TCE) to vinyl chloride (VC) under methanogenic
conditions (Vogel and McCarty, 1985).
17
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TABLE 7. TYPES OF ORGANIC POLLUTANT TRANSFORMATION ACCORDING TO THE
ROLE OF THE POLLUTANT IN MICROBIAL METABOLISM
Metabolic Role
Type of Metabolism
Outcome
Primary substrate (energy
source, carbon source,
electron donor)
Secondary substrate (could
be a carbon or energy source,
but present in too low
concentrations)
Electron acceptor
None
Primary substrate
utilization
Secondary substrate
utilization; occurring only in
the presence of a primary
substrate
Electron acceptor
utilization (respiration)
Cometabolism
Usually mineralization
(complete breakdown)
Usually mineralization
Partial reductive transforma-
tion (detoxification, or no
change in toxicity, or
activation)
Partial transformation
(detoxification, or no change
in toxicity, or activation)
In bacteria, incomplete chemical modifications of organic pollutants frequently result from a special
type of metabolism, called cometabolism. This phenomenon, which was first described (and
named) by Jensen (1963) and then fully analyzed by Horvath (1972), represents a partial chemical
change of the pollutant molecule which happens fortuitously and yields no benefit to the transform-
ing microorganism. The transformation of a cometabolic substrate happens while microorganisms
grow on their primary substrate. This transformation is due to the relaxed specificity of catabplic
enzymes with compromised active sites which will accommodate numerous structures, sometimes
only remotely related chemically. Aerobic transformation of chlorinated solvents, which is one of
the main topics of this report, is often cometabolic, such as TCE transformation by methanotrophs
(Little et al., 1988). It is important to emphasize, however, that cometabolism in a natural setting
may sometimes end as a complete degradation of the initially cometabolic substrate. Natural
habitats are inhabited by microbial communities which frequently practice interspecies transfer of
various metabolites; a product of cometabolic change, created fortuitously by the transformation-
initiating microorganism, may represent a useful resource for another microorganism(s) which will
catabolize it to inorganic products. Methanotrophs and heterotrophs together completely degrade
TCE, 1,2-dichloroethylene (1,2-DCE), VC, and other chlorinated compounds through such coop-
erative routes (Semprini et al., 1990,1991; Henry and Grbic-Galic, 1990, 1991a,b; Alvarez-
Cohen and McCarty, 1991a,b; Lanzarone and McCarty, 1990).
Under anaerobic conditions, halogenated compounds can undergo reductive dehalogenation
(Bouwer and McCarty, 1982; Suflita et al., 1982), a partial transformation which sometimes
detoxifies but in other cases activates the pollutant. This process is also included in cometabolic
reactions, with the active bacteria not getting any benefit from it. However, recent investigations
indicate that, in some cases, microorganisms may be utilizing the halogenated compounds as
electron acceptors, which implies generation of a useful form of energy when the halogenated
compound is coupled to a suitable electron donor. It has been suggested that PCBs may represent
such electron acceptors and that the microorganisms which reductively dechlorinate PCBs in
anoxic environments may be at an advantage relative to other members of the microflora (Brown et
18
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al., 1987). For some other chlorinated compounds, such as 3-chlorobenzoic acid, it has been
proven that the dehalogenating microbial communities indeed obtain energy from this dechlo-
rination (Dolfmg and Tiedje, 1987).
Acclimation Lag
Regardless of the type of microbial transformation of a pollutant, there is frequently an initial
period ("acclimation lag") during which no obvious changes of the pollutant occur. This period
may be due to various reasons (Linkfield et al., 1989). Sometimes the causes are in the indigenous
microbial communities. The starting biomass may be so low that no appreciable degradation can
happen until a critical biomass concentration is reached; or the total microflora may be abundant but
the specific active populations need to be enriched, or plasmids-coding for the active enzymes must
spread through the population. On other occasions, the pollutant must induce a requisite enzyme,
or a new enzyme needs to be synthesized, which will involve a genetic event such as mutation and
gene rearrangement. Sometimes, the reasons for the initial lag period may lie in the pollutants
themselves. The pollutants may be present in such low concentrations that they will not induce the
relevant enzymes or their chemical structure may be so unusual that they cannot interact with the
enzymic active sites. Finally, the environment will determine the duration of an acclimation period.
The lag can occur because preferential microbial substrates must be depleted first before the
degradation of a pollutant can start (in this case the pollutant is used as a primary substrate).
Sometimes, the temporarily inhibitory environmental conditions need to be changed; or, simply,
not all the necessary microbial nutrients are present, and microbial activity cannot be supported.
We can sometimes influence or change the environmental impacts on the acclimation lag; for
example, nutrients can be added to a contaminated habitat to stimulate microbial growth. If
pollutants are present in very low concentrations, suitable primary substrates can be injected into
the contaminated site to facilitate secondary substrate utilization. Further, if the desired physiologi-
cal group of microorganisms is not present, addition of acclimated or genetically modified
microorganisms to the site may be considered. Although this type of environmental biotechnology
is not yet practiced at large, it will be one of the most important technologies in the future.
However, if a pollutant is recalcitrant because of its xenobiotic chemical structure, as is the case
with some pesticides, little can be done once the contamination has occurred. The only viable
option in the case of recalcitrant pesticides is prevention (rather than cure): the use of biodegrad-
able alternatives with a similar range of activity.
ANAEROBIC TRANSFORMATIONS
The potential for anaerobic biological transformations of brominated and chlorinated (halogenated)
aliphatic compounds was demonstrated in 1981 (Bouwer et al., 1981). Subsequently, halogenated
aliphatic compounds in general have been found to transform under a variety of environmental
conditions in the absence of oxygen. In addition, research has indicated that some transformations
occur chemically (abiotic), as well as through the direct action of microorganisms (biotic). The
major abiotic and biotic transformation processes occurring in natural systems are summarized in
Table 8 (Vogel et al., 1987). The abiotic processes most frequently occurring under either aerobic
or anaerobic conditions are hydrolysis and dehydrohalogenation, while the anaerobic biotic
processes are generally reductions through hydrogenolysis or dihalo-elimination. The biotic
reductive processes appear to be occurring mostly through cometabolism. Here the compounds are
not used for energy or growth by the microorganisms, but are transformed fortuitously by
enzymes being used for other normal metabolic processes. Cometabolism generally occurs when
other chemicals are also present that can be used for energy and growth. This helps explain why
transformations occur at some sites but not at others.
19
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TABLE 8. TRANSFORMATIONS OF HALOGENTED ALIPHATIC COMPOUNDS
(after Vogeletal., 1987)
Reactions
Examples
I. Substitution
a. solvolysis, hydrolysis
RX + HgO—>• ROM + HX
b. other nucleophilic reactions
RX + NT —*> RN + X"
CH3CH2CH2Br
Br + MS"
CH CH CH OH + HBr
322
+ Br"
II. Oxidation
a. °< ~ hydroxylatton
C-X + H2O-*- -C-X + 2H++ 2e~
H
OH
b. epoxidation
v /*
2e
CH3CHd2 + H2O
CHCKXJI
2H+ + 2e~
2e~
III. Reduction
a. hydrogenolysis
RX + H++ 2e~-^> RH + X
CCI4+ H+ + 2e~_^. CHCI3+
b. dihalo-elimination
11 \ /
CCICCI
33 +
X X
c. coupling
2RX+2e~—> R-R +2X"
CCI2CCI2 + 2Cl'
2 CCI4 -i- 2e —^ CCI3CCI3 + 2CI
IV. Dehydrohalogenation
HX
,CCI2CH2 + HCI
20
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Figure 7 illustrates the anaerobic biotic and abiotic pathways that chlorinated aliphatic compounds
may undergo at contaminated sites. For example the chlorinated solvent TCA may be transformed
abiotically to form 1,1-DCE and acetic acid. The rates are relatively slow, with a half life for TCA
on the order of one year (Vogel et al., 1987; Cline and Delfino, 1989; Jeffers et al., 1989). Also,
under anaerobic conditions, TCA may be biologically transformed into 1,1-DCA, which can be
further reduced to CA. CA is relatively stable biologically; but abiotically it can be transformed
into acetate and chloride, thus rendering it relatively nontoxic. Thus, when TCA is discharged to
soil, a variety of abiotic and biotic transformation products may be found there in later years. As
another example, TCE can be reduced anaerobically to either cis- or trans-1,2-DCE, both of which
can further be transformed into vinyl chloride (VC). Recent research has indicated that VC can
even undergo reduction into ethylene (Freedman and Gossett, 1989), which is essentially
harmless.
The pathways outlined in Figure 7 suggest that it is possible to render harmless numerous chlori-
nated aliphatic compound under anaerobic conditions. Unfortunately, there are several problems
that hinder this potential approach to bioremediation. First, the biotic transformations generally
involve cometabolism such that other organic compounds must be present to serve as primary
substrates for organism growth. Second, the rates of anaerobic transformation are much greater
for the highly chlorinated compounds than for the less-chlorinated compounds, so that the less
chlorinated ones persist longer in the environment. Third, some of the anaerobic transformation
products are more hazardous than the parent compounds. Examples here are TCE transformation
to VC, and TCA transformation to 1,1-DCE. Fourth, reaction rates tend to be greater under highly
reducing conditions associated with methane formation than under the less reducing conditions
associated with denitrification (Bouwer and Wright 1988). The latter is the main anaerobic process
occurring when excess nitrates are present. Reductive transformation rates are somewhat interme-
diate between the two under conditions favoring sulfate reduction (sulfate, but no nitrate present).
Fifth, when proper environmental conditions are present, microorganisms that can bring about the
transformations through cometabolism must also be present.
cci3 cci3
CHCUCHCI CH2 =CCI2
cr
ANAEROBIC TRANSFORMATIONS OF HALOGENATED
ALIPHATIC COMPOUNDS
Figure 7. Chemical and biological transformation pathways of selected chlorinated aliphatic
compounds under anaerobic conditions. Arrows with "a" indicate chemical transfor-
mations; other arrows represent biological transformations (after Vogel, et al., 1987).
21
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Thus, one cannot count upon sufficiently high rates and complete transformation to harmless
products to occur in groundwater unless all the right conditions are present. On the other hand,
anaerobic transformation processes do frequently occur, converting chlorinated aliphatic
compounds into less chlorinated species that are more readily transformed by aerobic
microorganisms. It is for this reason, as well as to help understand the environmental fate of
compounds, that knowledge of anaerobic pathways is important.
AEROBIC MICROBIAL TRANSFORMATION OF Cl AND C2 CHLORINATED
ALIPHATIC HYDROCARBONS
Under aerobic conditions, halogenated aliphatic compounds with one or two carbons per molecule
can be transformed by three types of microbial enzymes: dehalogenases, hydrolytic dehalogenates,
and oxygenases. Dehalogenases, which require reduced glutathione as a cofactor, dehalogenate
the substrates by means of nucleophilic substitution (Stucki et al., 1981; Kohler-Staub and
Leisinger, 1985). The first product of this degradation pathway is an S-chloroalkyl-glutathione,
which is probably nonenzymatically converted to glutathione and an aldehyde (Stucki et al., 1981).
Hydrolytic dehalogenases hydrolyze their substrates, yielding alcohols (Goldman et al., 1968;
Keuning et al., 1985). Oxygenases use molecular oxygen as a reactant for the attack on the
halogenated compounds (Dalton, 1980; Hou, 1984; Nelson et al., 1987); the products could be
alcohols, aldehydes, or epoxides, depending on the structure of the compound. Numerous
halogenated short-chain aliphatic hydrocarbons have been demonstrated to undergo aerobic trans-
formation. However, compounds which have all the available valences ontheir carbon atoms
substituted by halogens, such as PCE or carbon tetrachloride, have never been shown to transform
through any other but reductive pathways (Vogel et al., 1987; Henson et al., 1988; Oldenhuis et
al., 1989). Generally, as the degree of halogenation increases, the likelihood of aerobic transfor-
mation decreases (Figure 8); the opposite is true for anaerobic (reductive) transformations (Vogel et
al., 1987).
Among Ci compounds, dichloromethane (DM) and chloroform (CF) have been found susceptible
to aerobic microbial transformation. DM can be completely mineralized under aerobic conditions
by sewage sludge microorganisms (Rittmann and McCarty, 1980a; Klecka, 1982) and by mixed
methanotrophic cultures enriched from soil (Henson et al., 1989). Pure cultures of the genera
Pseudomonas and Hyphomicrobium have been isolated which can grow on DM as the sole carbon
and energy source (Brunner and Leisinger, 1978; Brunner et al., 1980; Stucki et al., 1981; La Pat-
Polasko et al., 1984; Kohler-Staub et al., 1986). CF degradation has been reported for soil com-
munities consisting of methanotrophs and heterotrophs (Strand and Shippert, 1986; Alvarez-Cohen
and McCarty, 199 Ib). Pure cultures of methanotrophs, such as Methylosinus trichosporium
OB3b, partially transform (oxidize) DM and CF (Oldenhuis et al., 1989,1991).
Alkylhalides (haloalkanes), such as 1,2-dichloroethane (1,2-DCA), are frequently hydrolytically
dehalogenated (Stucki et al., 1983). Xanthobacter autotrophicus utilizes 1,2-DCA as sole carbon
source (Janssen et al., 1985). A hydrolytic dehalogenase isolated from this microorganism dehalo-
genates 1,2-dibromoethane (EDB) as well, but the bacterium cannot grow on this compound
(Keuning et al., 1985). An Acinetobacter sp., isolated from sewage sludge, was shown to dehalo-
genate and grow on ethylbromide (EB); EDB was dehalogenated, but not utilized for growth
(Janssen et al., 1987b). Complex communities consisting of methanotrophs and heterotrophs
which inhabit groundwater aquifers mineralize 1,2-DCA (Lanzarone and McCarty, 1990). A
Pseudomonas fluorescent strain isolated from water and soil contaminated by chlorinated aliphatic
hydrocarbons was shown to utilize 1,2-DCA, 1,1,2-trichloroethane (1,1,2-TCA), and TCE, but
not PCE or 1,1,1-TCA (Vandenbergh and Kunka, 1988). On the other hand, Methylosinus
trichosporium OB3b could transform not only 1,1-DCA and 1,2-DCA, but also 1,1,1-TCA
22
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1987,1988; Arciero et al., 1989; Wackett et al., 1989; Folsom et al., 1990; Harker and Kim,
1990; Vannelli et al., 1990). All of these microorganisms, except the genus Nitrosomonas, are
heterotrophs which grow on various organic substrates (e.g., aromatic hydrocarbons, phenols,
propane, etc.); Nitrosomonas is a chemolitotroph which derives energy from oxidation of
ammonia. All of them cometabolize chlorinated compounds such as TCE or 1,2-DCE while
growing on their respective growth substrates; the haloalkenes are only fortuitously transformed,
not utilized for growth. However, vinyl chloride seems to be an exception. It has been
demonstrated that a Mycobacterium strain isolated from soil contaminated by VC, could grow on
VC as sole carbon and energy source (Hartmans et al., 1985).
AEROBIC TRANSFORMATION AND DEGRADATION OF TCE
In 1985, Wilson and Wilson showed that TCE may be susceptible to aerobic degradation (by soil
microbial communities fed natural gas). Since then, much scientific research addressing this
phenomenon has been performed, including numerous laboratory experiments but also directed
field experiments (Roberts et al., 1990; Semprini et al., 1990,1991). So far, the groups of
bacteria capable of transforming TCE comprise methanotrophs (Fogel et al., 1986; Little et al.,
1988; Mayer et al., 1988; Tsien et al., 1989; Oldenhuis et al., 1989,1991; Henry and Grbic-Galic,
1990, 1991a,b; Alvarez-Cohen and McCarty, 1991a,b; Lanzarone and McCarty, 1990), propane
oxidizers (Wackett et al., 1989), ethylene oxidizers (Henry, 1991), toluene, phenol, or cresol oxi-
dizers (Nelson et al., 1986, 1987,1988; Wackett and Gibson, 1988; Folsom et al., 1990; Harker
and Kim, 1990), and ammonia oxidizers (Arciero et al., 1989; Vannelli et al., 1990). All of these
microorganisms have catabolic oxygenases that catalyze breakdown of their respective growth
substrates, but have nonspecific active sites which can accommodate TCE (and a variety of other
non-growth substrates) as well.
Thus TCE can be transformed (upon the induction of the oxygenase enzyme by its substrate) in the
presence of the microorganismal growth substrate (cometabolism). or in its absence (resting cells
transformation). However, TCE is not utilized by the bacteria as a carbon, energy, or electron
source; this transformation is only fortuitous. Based on the findings with methanotrophs (Little et
al., 1988; Henry and Grbic-Galic', 1991a), it can be concluded that TCE is most likely oxygenated
to TCE-epoxide. The epoxide is unstable and is quickly nonenzymatically rearranged in aqueous
solution to yield various products including carbon monoxide, formic acid, glyoxylic acid, and a
range of chlorinated acids (Miller and Guengerich, 1982). Recent findings with purified MMO
from Methylosinus trichosporium OB3b indicate that TCE-epoxide is indeed a product of TCE
oxygenation (Fox et al., 1990). In nature, where cooperation between the TCE oxidizers and other
bacteria (most prominently heterotrophs) occurs, TCE can be completely mineralized to carbon
dioxide, water, and chloride (Fogel et al., 1986; Henson et al., 1989; Roberts et al., 1989; Henry
and Grbic-Galic, 199 la).
Toluene, phenol, and cresol oxidizers, such as Pseudomonas putida or P. cepacia, express the
TCE transformation activity upon induction by their aromatic substrates (Nelson et al., 1988;
Folsom et al., 1990). These bacteria have a great potential for remediation of groundwater aquifers
which are contaminated by mixtures of gasoline or jet fuel (or other petroleum derivatives), and
chlorinated solvents, such as TCE, DCE, or VC. If the aromatic contaminants are not present,
however, bacterial growth substrates need to be injected into the site in order to stimulate the
transformation of chlorinated solvents. In this situation, methanotrophs become more attractive
agents of bioremediation: methane, their preferred substrate, is a nontoxic and inexpensive
chemical. Once methane and oxygen are injected into the site, methanotrophs (if present) will start
cometabolizing chlorinated solvents, as well as a great number of other contaminants (see below),
and the accompanying heterotrophs will mineralize their transformation products.
24
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Transformation and Degradation of TCE by Methanotrophs and Methanotrophic Communities
Methanotrophs grow on Q compounds as sole carbon and energy sources (Anthony, 1982).
Their catabolic oxygenases are methane monooxygenases (MMO) which incorporate one atom of
oxygen from the oxygen molecule into methane to yield methanol. This alcohol is further oxidized
via a series of dehydrogenation steps, through formaldehyde and formic acid, to CO2 which is the
final product of catabolism. MMO enzymes utilize molecular oxygen as a reactant, and require a
reduced electron carrier (e.g., NADH + H+) to reduce the remaining oxygen atom to water.
NADH + H+ is regenerated via dehydrogenation of catabolic intermediates. MMO enzymes have
relaxed substrate specificity, and will oxygenate many compounds which are not growth substrates
for methanotrophs. Such compounds include various alkanes, alkenes, ethers, alicycles,
aromatics, nitrogen heterocycles, and halogenated alkanes, alkenes, and aromatics (Colby et al.,
1977; Higgins et al., 1979; Stirling and Dalton, 1979; Stirling et al., 1979; Hou, 1984).
Two types of MMO have been suggested: a particulate (membrane-bound), and a soluble enzyme
(Dalton et al., 1984). The soluble MMO (purified from Methylosinus trichosporium OB3b and
Methylococcus capsulatus [bath]), which is produced under the conditions of copper limitation and
increased oxygen tension (Stanley et al., 1983), has been considered to have broader substrate
specificity (Colby et al., 1977; Burrows et al., 1984). It has been stated that only the soluble
MMO can transform TCE (Oldenhuis et al., 1989; Tsien et al., 1989). However, recent findings
indicate that the particulate MMO in some methanotrophs may be as effective in the transformation
of chlorinated solvents as the soluble MMO (Henry and Grbic-Galic, 1991a). Since the soluble
MMO is not constitutively expressed whereas the particulate MMO is, the latter methanotrophs
(Methylomonas sp.) have a significant potential for in-situ bioremediation.
On the basis of certain morphological and physiological differences, methanotrophs can be
classified as type I, type II, and Type X (Anthony, 1982). Types I and n have been shown to
oxidize TCE, presumably to an epoxide (Little et al., 1988; Oldenhuis et al., 1989; Tsien et al.,
1989; Henry and Grbic-Galic, 1990,1991a). Some of the suggested products of TCE-epoxide
hydrolysis, such as carbon monoxide, formic acid, glyoxylic acid, and chlorinated acids, have
indeed been detected in the methanotrophic culture fluid (Little et al., 1988; Henry and Grbic-
Galic, 1990,1991a,b). These products are quickly consumed by heterotrophic bacteria, so the
overall process brings about complete destruction of TCE (Henson et al., 1989; Alvarez-Cohen
and McCarty, 1991a; Henry and Grbic-Galic, 1991a). In groundwater in situ, such degradations
have been demonstrated at a moderate efficiency for TCE and cis-l,2-DCE, and at a great
efficiency for trans-l,2-DCE and VC (Semprini et al., 1990,1991).
Significant Findings Concerning the Fortuitous Transformation of TCE by Methanotrophs
Like any other microbial process affecting an environmental pollutant, methanotrophic transforma-
tion of TCE has certain peculiarities and also requirements which need to be satisfied in order for
the process to function. Furthermore, the transformation is susceptible to environmental influ-
ences. In the course of the laboratory investigations of TCE transformation at Stanford University,
under the auspices of the Moffett Field project (Roberts et al., 1989), a number of interesting
phenomena were observed. Several of these phenomena were confirmed through parallel work of
researchers at other institutions. The laboratory batch cultivation techniques used and the most
important of the findings are described below.
Batch Cultivation Techniques~For microcosm work, glass columns (40x2 cm) were packed with
sandy aquifer material. The columns were batch-fed methane and oxygen (or hydrogen peroxide)
dissolved in groundwater. Cold or 14C-labeled TCE was added to the feed after the microbial
communities had become established in the columns. The sequential exchanges of the pore water
25
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were performed at intervals of 3 days, one week, or two weeks. The effluent samples were
analyzed for methane concentration (gas chromatography, GC), dissolved oxygen (dissolved
oxygen probe), hydrogen peroxide concentration (titration with eerie sulfate), number of bacteria
(acridine-orange direct counting—Ghiorse and Balkwill, 1983), TCE concentration, and other 14C-
labeled species (GC, scintillation counting). A total of six columns were operated; two of them
were live controls that received no methane (Mayer et al., 1988).
The suspended mixed cultures were enriched from aquifer material, groundwater, or laboratory
column effluent. The enrichments were achieved by inoculating continuously stirred reactors
containing defined mineral medium. The reactors were incubated under a continuous flow of
approximately 25% methane in air, at room temperature (21-23°C). The transfers to new reactors
were repeated 5-10 times, until stable methanotrophic consortia were obtained (as determined by
micro- and macroscopic observations). Pure methanotrophic cultures were isolated from these
mixed cultures on agarose plates with mineral medium. The plates were incubated in desiccators
filled with 25% methane in air, at room temperature. Purity was determined through repeated
colony isolation and observation of constant macro- and micromorphology (Henry, 1991). The
isolates were tested for the lack of growth on multicarbon substrate in liquid and solid media, and
for the type of biosynthetic enzymes (Large and Quayle, 1963; Dahl et al., 1972). Both mixed and
pure cultures were examined by light microscopy (including staining procedures for the determina-
tion of membrane structure and presence of internal storage granules), and scanning and transmis-
sion electron microscopy (Henry and Grbic-Galic, 1990).
For TCE transformation experiments, the cultures were first raised to mid-logarithmic growth
phase in defined mineral medium in continuously stirred reactors under a continuous stream of 30-
35% methane in air, at room temperature. The pure culture received a mixture of vitamins for
growth in liquid medium (Henry, 1991). The inocula from the reactors were then transferred to
250-ml screw-cap bottles containing 150-ml headspace. The bottles were incubated upside-down
on a rotary shaker in a 21°C environmental chamber, for resting-cell transformation studies.
Formate, when added to some of the cultures as an electron donor, was provided as 2 mM sodium
formate. The cultures were appropriately diluted (to prevent mass-transfer limitations) with either
phosphate buffer or the same mineral medium in which they had been grown. Cell biomass was
determined as dry weight. TCE concentration and TCE transformation products were evaluated in
the bottle headspace and in the aqueous fraction by GC, gas partitioning, ion chromatography, and
reduction gas detector (for carbon monoxide). Scintillation counting was used in the experiments
with 14C-labeled TCE (Henry and Grbic-Galic, 1990). Autoclaved controls were used in all the
experiments.
Competitive Inhibition-MMO is the enzyme which is implied in oxidation of both methane and
TCE (or other organic pollutants transformed by methanotrophs). Therefore, methane and TCE
will compete for the same active site on the MMO. Depending on the affinities of the enzyme for
methane and TCE (Ks), which vary in different methanotrophs (Oldenhuis et al., 1989; Tsien et
al., 1989; Alvarez-Cohen and McCarty, 1991a; Henry and Grbic-Galic, 1991a), and on the con-
centrations of methane and TCE, methane will interfere more or less efficiently with TCE transfor-
mation (and vice versa) if both compounds are simultaneously present.
The phenomenon of competitive inhibition by methane in cometabolism of TCE has been shown
for all methanotrophic systems examined so far. This inhibition can be relaxed if the methane
concentration is kept low (Henry and Grbic-Gali6,1990), or if the resting-cells transformation of
TCE (in the absence of methane) is employed (Oldenhuis et al., 1989; Alvarez-Cohen and
McCarty, 1991a; Henry and Grbic-Galic, 1991a). For example, resting cells of Methylomonas sp.
strain MM2 (Henry and Grbic-Galic, 1991a) transform TCE at low concentrations (30-60 |ig I'1),
in the presence of 2 mM formate as an electron donor, at a rate as high as 2.3±0.05 1 rag'1 d'1
26
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(pseudo-first-order rate constant, k' = k/Ks) at 21°C. The competitive inhibition of TCE
transformation by methane has to be taken into account when the process is considered for
application (Semprini et al., 1991).
The Importance of Endogenous or Exogenous Electron Donors—In the presence of methane,
NADH+H+ is regenerated through dehydrogenations of methanol, formaldehyde, and formate. In
the absence of methane, it is not possible to renew the reduced electron carriers necessary for the
oxygenation of TCE. However, some methanotrophs will continue transforming TCE in the
absence of methane (resting cells transformation) for extended periods of time (longer than 24 hr,
Henry and Grbic-Gali£, 1991a; see Figure 9.B), whereas others will lose the transformation
capability within several hours (Alvarez-Cohen and McCarty, 199la). The former group has been
shown to form intracellular lipid storage inclusions while growing on methane (Henry and Grbic-
Galic, 1990,1991a). This storage can be used as a source of electrons and protons for the
oxygenation reactions in resting cells.
10 20 30 40 50 60
TIME, hourt without methane
20 40 60
TIME, houra without nwthMM
Figure 9. The influence of formate addition on TCE transformation rates in the absence of
methane, in the pure culture Methylomonas sp. MM2 (A) and mixed culture MM1 (B),
both derived from the Moffett Field groundwater aquifer. MM1 contains methano-
trophs with lipid storage inclusions, and can transform TCE for a prolonged period of
time in the absence of methane, but it does not react to formate addition (B).
Methylomonas sp. MM2, on the other hand, does not have internal storage granules,
but its resting-cell transformation of TCE can be prolonged by formate addition (A).
"None" = subcultures not amended with formate. "None+formate" = subcultures
incubated without formate, but receiving 2 mM formate at the time of TCE addition.
"Formate" = subcultures incubated with formate, and receiving 2 mM formate at the
time of TCE addition (from Henry and Grbic-Galic, 1991a).
27
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The methanotrophs which do not have the endogenous electron donor storage can be aided in their
resting-stage fortuitous transformations if an external electron donor is added, such as formate
(Oldenhuis et al., 1989; Semprini et al., 1991; Alvarez-Cohen and McCarty, 1991a; Henry and
Grbic-Galic, 1991a) or methanol (Mayer and Grbic-Gali6,1989; Semprini et al., 1991). Figure 9
shows the behavior of a pure methanotrophic culture, Methylomonas sp. MM2, and a mixed
methanotrophic culture, MM1, both derived from the Moffett Field groundwater aquifer (Henry
and Grbic-Galic 1990), in the absence of methane but with the formate addition. Methylomonas
sp. MM2 (Figure 9.A) does not contain internal storage granules, whereas the mixed culture MM1
(Figure 9.B) does. As the result, formate increases the TCE transformation rates in Methylomonas
sp. MM2 under starvation conditions, while MM1 remains insensitive to the formate addition.
The methanotrophs containing storage granules which represent an internal electron donor source
may be of a great significance for in-situ treatment of contaminated aquifers. Such organisms
would be capable of prolonged contaminant transformation in the absence of methane, thus
circumventing competitive inhibition, and would not require the addition of alternate reductants.
Although formate or methanol can be used as alternate electron sources by some methanotrophs,
they are also utilized by heterotrophs; the addition of these compounds to the subsurface could thus
shift the delicate balance of the active microbial community.
Product Inhibition-Carbon monoxide (CO) is one of the products of TCE transformation by
methanotrophs: Methylomonas sp. MM2 was shown to transform approximately 9 mol% of the
TCE to CO, which was subsequently oxidized (Henry and Grbic-Galic, 1991b; see Figure 10). It
is known that CO can be oxidized by methanotrophs, and that the oxidation is catalyzed by MMO
(Ferenci, 1974; Hubley et al., 1974; Ferenci et al., 1975; Henry and Grbic-Galic, 1991b). Conse-
quently, the CO will compete for the MMO active site; and it will scavenge reduced electron/proton
donors (Henry and Grbic-Galic, 199Ib, have shown that the CO oxidation rate doubles upon
addition of formate as an electron source). As the result, a competitive inhibition of TCE transfor-
mation by CO occurs (Figure 11). Our experiments with Methylomonas sp. MM2 indicated that
CO is a much more powerful competitive inhibitor of TCE transformation than is methane (Henry
and Grbic-Galic, 1991b): the measured inhibition constant (Ki) for CO inhibition of TCE
oxidation was 4.2 (jM, and that for methane inhibition of TCE oxidation, 116 (iM. The CO itself
was not toxic to methanotrophs, but it interfered with methane oxidation by competing for the same
resources. The influence of CO may be significant in treatment of chlorinated solvents by pure
cultures of methanotrophs, and will depend on the affinities of the microorganisms for the
chlorinated solvent and CO, respectively, on the amount of CO produced, and the availability of
the reductant. However, this effect has not been demonstrated in mixed microbial communities
(Henry and Grbic-Galie, 1991b).
Chemistry of the Medium-The composition of the nutrient medium which is used to grow
methanotrophs can have a profound influence on subsequent TCE transformation. For example,
the presence or absence of copper will significantly influence the formation of soluble MMO in
methanotrophs which possess this type of enzyme, and consequently affect TCE transformation
(Oldenhuis et al., 1989). In the presence of copper, the soluble MMO is not formed, and the
subsequent TCE transformation becomes negligible.
Furthermore, the presence or absence of a metal chelator, such as EDTA, in the growth medium
can result in a change in affinity of the methanotrophic enzymes (Ks) for TCE. It can also result in
more than an order of magnitude difference in the TCE oxidation rate (k/Ks) under resting cell
conditions (Henry and Grbic-Galic, 1990; see Table 9). EDTA can chelate a variety of metal ions,
such as calcium, magnesium, copper, iron, and various trace metals (Bridson and Brecker, 1970;
Stumm and Morgan, 1981). Some of these metals become more available through chelation,
whereas others are rendered unavailable (O'Sullivan, 1969). The most important factors
28
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0.4
Abiotic control (TCE)
-0.5
0.5 1.5 2.5
Time, hours
0.03
-0.5
2.5 3.5
Time, hours
4.5
5.5
6.5
Figure 10. Production (from TCE) and subsequent oxidation of CO by Methylomonas sp. MM2
[(B) is an expanded view of the CO data from (A)]. Cell density: 0.11 g cells (dry
weight) I'1. Abiotic control: sterile mineral medium and TCE. Live control: a
subculture of Methylomonas sp. MM2, without TCE (from Henry and Grbi6-Galic,
1991b).
29
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0.03
0.02
o
0.01 -
0.00
2 4
Time, hours
Figure 1 1. Competitive inhibition of TCE oxidation in Methylomonas sp. MM2 by CO. Cell
densities: 0.18 to 0.21 g cells (dry weight) H. The cultures were amended with 2
mM formate. Symbols: (+) no CO; (A) 6.7 jiM CO; (A) 25.6 ^iM CO; (•) 53.5
CO; (Q) 57.3 M.M CO; (O) control (sterile mineral medium). From Henry and Grbic-
Galic, 1991b.
in this complex process are the stability constants of the metal-EDTA complexes, and the
concentrations of metals in the medium. EDTA potentially could affect the methanotrophic
performance in various ways, and the specific mechanism has not been elucidated yet, but there are
indications that copper, calcium, or magnesium availability may be crucial (Henry and Grbic-Galic,
1990). These results stress the significance of water chemistry in the TCE transformation process
in situ: the composition of solutes in groundwater will substantially affect the transformation of the
contaminant.
Particulate Versus Soluble MMO-- Although it is commonly believed that only the soluble form of
MMO (produced by Methylosinus trichosporiwn OB3b and Methylococcus capsulatus [bath]) is
capable of catalyzing TCE transformation (Oldenhuis et al., 1989; Tsien et al., 1989), our recent
work suggests that the paniculate form of MMO in some methanotrophs may be at least as efficient
(Henry and Grbi6-Gali6, 1990, 199 la). Methylomonas sp. MM2, isolated from Moffett Field
groundwater aquifer, has never been shown to produce the soluble MMO (Henry and Grbic-Galic,
1991a), and it is relatively insensitive to variations in copper concentration. The pseudo-first-order
rate constant (k/Ks) for TCE transformation in Methylomonas sp. MM2 (1.6 ml mg"1 mur1) is
comparable to that of Methylosinus trichosporium OB3b (2.14 ml mg-1 min'1).
30
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TABLE 9. KINETIC PARAMETERS FOR TCE TRANSFORMATION BY THE PURE
CULTURE METHYLOMONAS SP. MM2 GROWN IN THE PRESENCE OR ABSENCE OF A
METAL CHELATOR, EDTA
(from Henry and Grbi6-Galic. 1990)
Mineral Medium kCday1)0 KsCmgH)6 k/Ks
(1 mg-1 day1) c
Whittenbury^
Whittenbury, no EDTA
0.29
0.046
0.51
1.35
0.57
0,033
a k = maximum specific substrate utilization rate.
b Ks = half saturation coefficient.
c k/Ks = second order (pseudo-first-order) rate coefficient.
d Whittenbury = mineral medium formulation after Whittenbury et al. (1970).
The paniculate form of the enzyme could be very significant in environmental applications, because
the expression of this enzyme does not require the conditions of rigorous copper limitation. The
environmental strains with the constitutively expressed paniculate MMO which has a broad sub-
strate specificity, such as Methylomonas sp. MM2 (Henry and Grbic-Galic, 1990), are ideally
suited for in-situ treatment of contaminated groundwater aquifers.
TCE Oxidation Toxicity—TCE in high concentrations (e.g., 50 mg I'1) has been shown to inhibit
methane utilization and TCE transformation (Oldenhuis et al., 1989). The results of our research
show that some of the products of TCE oxidation by methanotrophs are in fact toxic to these bac-
teria (Alvarez-Cohen and McCarty, 1991a; Henry and Grbic-Gali6,1991a), even when the TCE
concentrations are considerably lower (below 10 mg I'1; Henry and Grbi6-Gali6,1991a). Similar
findings have been reported for Pseudomonasputida Fl, which transforms TCE upon stimulation
with toluene (Wackett and Householder, 1989). Previous work with mammalian liver enzymes
had shown that TCE epoxide and its products interacted with proteins and other macromolecules in
mammalian cells, causing damage (Bolt and Filser, 1977; Bergman, 1983).
Our results indicate that the 14C from 14C-labeled TCE becomes incorporated into the cells of
Methylomonas sp. MM2, although it is not likely that the products of hydrolysis of TCE-epoxide
could be utilized by the pure methanotrophic culture (Henry and Grbi6-Gali6, 199la). Further-
more, the methanotrophic cells which had transformed TCE have great difficulties in subsequent
utilization of their growth substrate, methane (Figure 12), and the overall active cell numbers are
reduced by more than an order of magnitude. The oxygenase enzymes in Methylomonas sp. MM2
and Pseudomonas putida Fl seem to activate TCE in a similar fashion as mammalian enzymes, and
the effect seems to be comparable: the inactivation of the transforming cells.
This phenomenon should be a very important consideration for application of methanotrophic
processes: it must be kept in mind that TCE oxygenation inactivates methanotrophs and that this
effect will be more pronounced at higher TCE concentrations and lower cell densities.
31
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"TCE" cells with
"none'r medium
40
TIME, hours
Figure 12. The effect of TCE oxidation on subsequent methane utilization by Methylomonas sp.
MM2. Culture medium from the subcultures incubated with 6 mg of TCE per liter
("TCE" subcultures) was switched with that from the subcultures incubated without
TCE ("None" subcultures). Only the "TCE" subcultures amended with "None" super-
natant were affected, indicating that the TCE transformation itself was toxic to the
methanotroph. The aqueous intermediates of TCE transformation, on the other hand,
were not inhibitory for methane utilization (after Henry and Grbic-Galic, 1991a).
The Influence of Hydrogen Peroxide-Hydrogen peroxide is frequently used as a substitute for
oxygen in groundwater remediation (Lee et al., 1988). However, our experiments with column
microcosms consisting of saturated aquifer material from the Moffett Field site indicate that H2Q2
may be inhibitory to methanotrophs (Mayer et al., 1988). Heterotrophs are less affected; however,
since in this particular case the methanotrophs are the initiating microorganisms in the TCE degra-
dation chain, the TCE-degrading activity is lost. The field experiments in the aquifer itself did not
show an improvement in TCE transformation efficiency upon addition of H2C>2 either (Semprini et
al., 1990). This should be another important consideration for the application of methanotrophs in
in-situ treatment of chlorinated solvents.
Important Considerations—Summary
Table 10 summarizes the most important factors to be considered in methanotrophic in-situ
treatment of chlorinated solvent contamination.
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TABLE 10. SUMMARY OF THE IMPORTANT FACTORS FOR IN-SITU TREATMENT OF
CHLORINATED SOLVENTS BY METHANOTROPfflC COMMUNITIES
Methane concentration (competitive inhibition of chlorinated solvent
transformation)
Transformation in the absence of methane (resting-cells) — alternate pulses of
CH4 and oxygen
Oxygen concentration
Influence of hydrogen peroxide as an oxygen substitute
Chlorinated solvent concentration (inhibition of microorganisms)
External (formate) or internal (lipid storage granules) electron donors in
resting-cells transformation
Particulate vs. soluble methane monooxygenase (paniculate seems to be less
susceptible to external influences, such as medium chemistry)
Robustness of microorganisms vs. high transformation rates (in a
subsurface environment, robustness may be far more important)
Groundwater chemistry (influence on microbial growth and transformation
capabilities)
Chlorinated solvent oxidation toxicity (reduction in number of viable
microorganisms)
Product inhibition of chlorinated solvent transformation (percent conversion of the
parent compound to a particular product, and the affinity of methanotrophs for the
product oxidation)
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SECTION 4
PROCESS MODELS
INTRODUCTION
Models used in simulating biotransformation processes in porous media are useful tools in design-
ing in-situ remediation systems. The models may be used in evaluating different remediation alter-
natives, determining effective means of implementing remediations, and providing estimates on
clean-up times, chemical demands, cost for pumping, etc. This section will present kinetic models
for microbial processes including cometabolic transformations. The incorporation of these micro-
bial process models into the advective-dispersive transport equation will also be introduced.
Models are still being developed for cometabolic transformation processes, so it is premature to
present details of different models that have been developed. A model that was used in evaluating
the results of a field demonstration of in-situ biodegradation of chlorinated aliphatics (Section 5)
will be discussed. In Section 6 the model will be used as a tool for evaluating the bioremediation
potential of the methanotrophic process at a Superfund site in Michigan.
The enhanced in-situ biotransformation approach requires creating a biologically reactive zone in
the subsurface. Conceptual models for in-situ treatment systems are given by Lee et al. (1988) and
McCarty (1985). The conceptual model shown in Figure 13 presents several forms of biological
treatment: 1) surface treatment, 2) a well bore reactor, and 3) in-situ treatment of the contaminated
aquifer. The biological treatment of the halogenated aliphatics may be accomplished by one or a
combination of these forms of treatment. There is as yet insufficient experience with biological
treatment of chlorinated aliphatics to determine the appropriate conditions for in-situ treatment, and
to judge categorically when in-situ or above ground treatment should be applied. Models that
include the key microbial and transport processes can be useful in determining the important factors
to consider in selecting a treatment process and whether there may be an advantage to using one
treatment process or a combination of processes.
Bioremediation depends on a combination of complex processes including advective and dispersive
transport, sorption, microbial growth, utilization of electron donors and acceptors, and for cometa-
bolic transformations, the transformation of the contaminant itself. Modeling of these coupled
processes is the only means for evaluating their interactions and for determining how bioremedia-
tion will be affected by the combination of the different processes.
The main processes that need to be considered are advection, dispersion, sorption and biodegrada-
tion (Figure 14). Transport processes include advection (idealized as plug flow in Figure 14) and
dispersion which acts to spread the contaminants compared to idealized plug flow. Sorption
partitions the contaminants between the aqueous and solid phases. Biodegradation acts as a sink
term that removes contaminants, and is usually considered to occur in the aqueous phase. Figure
14 shows conceptually the concentration histories of contaminants at an observation location that
result from their continuous injection into the subsurface. Concentration histories are shown for
contaminants that do and do not sorb, and are and are not biodegraded. Sorption onto the aquifer
34
-------
SUBSTRATE
AND
NUTRIENTS
EXTRACTION
WELL
*y/s
1 ABOVE GROUND
* BIOREACTOR
BYPASS
v/ v/x y// y//
•*— AQUIFER
-— BIOREACTOR
?// y//
W
oU
l< ^
NU
> y?/'
SUBSTRATE
AND
NUTRIENTS
WELL-CASING
BIOREACTOR
Figure 13. Conceptual diagram of above-ground well casing and aquifer (in-situ) bioreactor (from
McCarty, 1985).
EXPECTED RESPONSES TO 4 STEP CHANGE
IN CONCENTRATION
123456
TIME RELATIVE TO MEAN RESIDENCE
TIME OF WATER , I / T
Ideal Plug F!ow I? = - ' 3dT
Dispersion (Pick's Law) ^~ = DK TT
Sorplion (Retardation)
Biodegradation (Sink)
82c a
biodegradahon
sink term
Figure 14. Conceptual model of transport and biodegradation processes that must be considered
(from McCarty et al., 1981).
35
-------
solids results in slower transport through the aquifer and longer transport times, both in the pres-
ence and absence of biotransformation. Biodegraded contaminants initially break through like the
nondegraded contaminants. As time proceeds, the contaminant concentrations decrease as
microbes grow that can degrade the contaminants. Methane as a growth substrate would be
expected to behave like the nonsorbing, biodegraded contaminant The chlorinated aliphatics
would be expected to behave as contaminants that are sorbed and biodegraded.
As indicated in Figure 14, many processes must be incorporated in simulating the degradation of
contaminants in the subsurface. To illustrate the processes, one-dimensional representations of the
mathematical terms are shown. Here sorption is represented by a retardation model as described in
Section 2. The biological sink term (A,) represents the rate of transformation of the contaminants.
For cometabolic transformations, this will be shown to be a complicated term, dependent on many
factors.
MICROBIAL PROCESSES
Monod kinetics and variations of Monod kinetics are most commonly used in the formulation of
rate expressions for microbial processes, including microbial growth, utilization of the electron
donor and acceptors, and for the cometabolic transformations of contaminants. Monod kinetics are
derived from Michaelis-Menten enzyme kinetics (Figure 15). Here the reaction velocity (V) is
plotted versus substrate concentration (S). At high substrate concentrations (S » Km), the rate of
reaction (V) approaches a maximum rate (Vmax) that is independent of substrate concentration;
while at low substrate concentrations (S « Km), the rate of reaction, V = VmaxS/Km, is first order
with respect to substrate concentration. This first-order reaction rate dependence has implications
on transformation rates at low concentrations, as the reaction rates become progressively slower
with decreasing substrate concentration.
I
Substrate concentration [S]
V =
Michaelis-Menten (19131
V = rate = (mol/diter-sec)
Vmll = maximum utilization rale (mol/liter-sec)
Kn = Michaelis constant (mol/littrl
v 5
S « k, V = V'
Km
S »» k, V = Vnl,
Figure 15. Michaelis-Menten enzyme kinetics (from Stryer, 1981).
36
-------
The Monod equation is commonly used for the rate of primary substrate utilization:
Cp
dt
(6)
where X = concentration of microbes (mg/1), k = maximum substrate utilization rate (day1), KSD
= substrate saturation constant (mg/1), and CD = electron donor concentration (mg/1).
The rate of substrate utilization (electron donor) is proportional to the microbial concentration,
assuming that the enzyme concentration is also proportional to the microbial concentration (X).
The substrate concentration term is the same as the enzyme kinetic model. For utilization by
methanotrophs, CD is the methane concentration. When the utilization of the primary substrate is
dependent on the presence of oxygen, as is the case for methanotrophic bacteria, a dual form of
Monod kinetics is often used:
dCp _ _._. / Cp y CA
dt ~ KA KSD +
where CA = oxygen concentration (mg/1) and KSA = contaminant saturation constant for oxygen
(mg/1). Thus, the rate of substrate utilization can be limited by both the methane or oxygen concen-
tration.
Kinetic models for the cometabolic transformation of chlorinated aliphatics are currently under
investigation. One model that appears to be appropriate for the transformation of chlorinated
aliphatics by methanotrophs is a form of the Monod model that includes terms for competitive
inhibition. Competition between the growth substrate and the non-growth substrates for the active
site of MMO enzyme can affect the rate of transformation of the non growth substrate as discussed
in Section 3. Hou et al. (1979) and Patel et al. (1982) found methane to inhibit the rate of transfor-
mation of other hydrocarbons. Recent studies have shown that methane can inhibit the transforma-
tion rates of chlorinated aliphatics (Strand et al. 1990, Lanzarone and McCarty, 1990, Oldenhuis et
al., 1991, Semprini et al., 1991). Equation 8 is a dual Monod kinetic model that has been adapted
for competitive inhibition (Semprini et al., 1991).
dC2
dt
VI,-
— .A.K2
r c2 i
KS2
+ C2H
Kj.
where X = concentration of contaminant degrading microbes (mg/1), k2 = maximum contaminant
transformation rate (day1), KI = inhibition constant (-), K$2 = contaminant saturation constant
(mg/1), C2 = concentration of non-growth contaminant (mg/1), Q = inhibitor concentration (mg/1),
and CA = oxygen concentration (mg/1).
For methanotrophs, methane inhibits the rate of contaminant biodegradation; thus Q in Eq. 8 is
the methane concentration CD- According to this model, the rate of contaminant transformation
decreases with increasing methane concentration. The effect is more pronounced the lower the
value of Kj. A second Monod factor is included for the electron acceptor, since the presence of
dissolved oxygen is required for contaminant transformation.
37
-------
Equation 8 illustrates that many factors can influence the rate of cometabolic transformation by
methanotrophs. The goal of enhanced in-situ bioremediation is to increase the biomass concen-
tration, X, to enhance the rate of the cometabolic transformation. For methanotrophs this requires
the addition of methane as a primary substrate for growth. As indicated in Eq. 8, the addition of
high methane concentrations to increase the methanotrophic biomass in the aquifer could adversely
affect transformation rates due to competitive inhibition by methane.
The rate of the cometabolic transformation by methanotrophs, both spatially and temporally,
depends on the microbial mass, methane and oxygen concentrations (Eq. 8); consequently, in
mathematical model computations all dependent variables must be calculated simultaneously. The
microbial processes that must be considered include methanotrophic growth, the utilization of
methane as a primary substrate (electron donor) and of oxygen as an electron acceptor, and the
cometabolic transformation of the halogenated aliphatic.
The rate of microbial growth and decay of microbial populations can be represented by:
dt
Cp \ / CA >> . „ / CA
A \
+ CAJ
(9)
where Y = yield coefficient (mg bacteria/mg substrate) and b = decay coefficient (day1). The dual
Monod kinetic formulation shows that the growth of methanotrophic bacteria can be limited by the
absence of methane or oxygen. In this equation, the rate of decay also depends on the oxygen
concentration.
Equation 7 gives the dual Monod expression for electron donor utilization. The dual Monod
expression is also used for the utilization of the electron acceptor:
where F = stoichiometric factor (mg acceptor/mg donor), dc = cell decay oxygen demand (mg
acceptor/mg cells), and fj = degradable cell fraction. For methanotrophs, the stoichiometric factor
represents the amount of oxygen consumed for the amount of methane utilized. The rate of oxygen
consumption due to biomass decay is also included in Eq. 10.
Biodegradation models based on the variations of the Monod equation presented above have been
developed for simulating the utilization of organics and oxygen in porous media. A review of how
these models are incorporated into transport equations is given by Baveye and Valocchi (1989).
Tables 1 1 and 12 give citations and brief descriptions of some of the models that have been devel-
oped. The simplified models given in Table 1 1 are for shallow bipfilms, and the more complicated
biofilm models are listed in Table 12. The model developed for simulation of the results of the
Moffett Field study, which will be discussed later, used the simplified model of shallow biofilm
kinetics.
COUPLING WITH TRANSPORT PROCESSES
The transport of the microbes, electron donors, electron acceptors, and halogenated contaminants
is an important component in modeling subsurface biotransformation processes. In the microbial
kinetic models, source and sink terms are added to the advective-dispersive transport equation.
Details of this approach are given in the papers cited in Tables 1 1 and 12.
38
-------
TABLE 11. SELECTED SHALLOW BIOFILM MODELS FOR MICROBIAL
TRANSFORMATION IN POROUS MEDIA
Borden and Bedient (1986) - Developed model for the instantaneous reaction of a primary
substrate with oxygen.
Borden et al. (1986) - Applied model of Borden and Bedient (1986) to aquifer
contaminated with fuels.
Corapcioglu and Haridus (1985) - Developed a model that considers biofouling as a result
of biogrowth.
Kindred and Celia (1989) - Developed model that includes aerobic and anaerobic reactions
plus cometabolic transformations.
Molz et al. (1986) -- Developed non-steady-state model for growth of micro-colonies from
the utilization of electron donor and acceptor.
Semprini and McCarty (1989) ~ Results of a field study simulated using a model that
considers co-metabolic transformations and rate-limited sorption-desorption.
Srinavasan and Mercer (1988) - Developed model that considers different sorption
processes along with primary substrate utilization.
TABLE 12. SELECTED BIOFILM MODELS FOR MICROBIAL TRANSFORMATIONS
Bouwer and McCarty (1984) -- Applied model of Rittmann and McCarty (1980b) to tracer
organic biotransformation in the subsurface.
Bouwer and McCarty (1985) - Used model of Rittmann and McCarty (1980b) for
transformation in laboratory columns of trace halogenated organics as secondary
substrates.
Bouwer and Wright (1988) — Modeled the transformation of chlorinated aliphatics in
laboratory column operated under different electron acceptor conditions.
Bryers (1988) - Modeled biofilm accumulation of mixed cultures.
Kissel et al. (1987) ~ Modeling mixed culture biofilms with different electron acceptors
used in the biofilm.
Rittmann and McCarty (1980b) ~ Analytical model developedfor steady-state biofilms and
secondary-substrate utilization.
Rittmann et al. (1980) ~ Applied model of Rittmann and McCarty (1980b) to results of a
field experiment.
Rittmann et al. (1988) — Modeled transformation of chlorinated aliphatics in a laboratory
column operated under denitrifying conditions.
Speitel et al. (1987) ~ Biofilm modeling of GAC columns including sorption and radial
diffusion into the activated carbon.
39
-------
Different assumptions are made in models of how to represent the microbial mass in these models
as discussed by Baveye and Valocchi (1989). Most models assume that the microbial mass is
attached to the porous media and is immobile. Most models also assume that biodegradation
occurs only in the aqueous phase. The sorbed contaminants must desorb in order to be trans-
formed, as indicated in Figure 16 (McCarty, 1988).
Partitioning of the contaminants between the aqueous and solid phases must therefore be consid-
ered. Sorption can cause a large fraction of the contaminant mass to be partitioned onto the aquifer
solids. Sorption also lowers the aqueous concentration, thus reducing the rate of reaction (Eq. 9).
Biodegradation in the aqueous phase reduces the aqueous concentration, creating a driving force
for contaminant desorption from the aquifer solids. The desorption is a physical process that is
being driven by a biological process.
IMASS TRANSFER CONSIDERATIONS!
Figure 16. Mass transfer considerations (from McCarty, 1988).
The sorption and desorption of a contaminant can be modeled as an equilibrium or rate-limited
process (Section 2). Laboratory studies, presented in Section 2, indicate that sorption-desorption
of aquifer solids can be rate-limited. Thus, models should consider sorption-desorption as a rate-
limited process, when appropriate.
A model must also be selected for the partitioning of the contaminants onto the aquifer solids.
As discussed in Section 2, the simplest model, and probably the most adequate for halogenated
aliphatics, is a linear and reversible sorption model (linear isotherm), with the equilibrium sorbed-
phase concentration given by:
C =
(11)
where IQ is the partition coefficient (I/kg) and C is the sorbed-phase concentration (mg/kg).
This simple sorption model is easily incorporated into a model that includes rate-limited transfer
between the solid and aqueous phases. One of the simplest forms of a rate-limited sorption model
that is easily incorporated into the advection-dispersion equation is the first-order linear nonequilib-
rium model, also known as the chemical kinetic model:
40
-------
= a (Kd C - G)
(12)
where a is the rate coefficient for mass transfer between phases. The driving force for mass
transfer is the concentration difference between the aqueous and sorbed phase, where the aqueous-
phase concentration is expressed in terms of the equilibrium sorbed-phase concentration using the
linear sorption model. Desorption occurs when the expressed equilibrium-sorbed concentration is
less than the actual concentration on the solids, and sorption occurs when the opposite conditions
hold true. This simple kinetic model represents a reasonable approximation of more complex sorp-
tion models that include diffusive transfer between mobile and immobile zones (van Genuchten,
1985). This kinetic model was used in simulating the Moffett Field chlorinated organic responses
without biotransformation (Harmon et al., 1990) and with biotransformation (Roberts et al., 1989;
Semprini and McCarty, 1992).
Equation 12 is incorporated into the advection-dispersion equation as a source sink term:
8C ,. 32C 3C pb fv „ ^
= ^- a(KdC-c)
Upon de_sorption (Kd C < C), it acts as a source term to the aqueous phase, while for sorption
> C), it acts as a sink term. Equation 12 must also be solved along with Eq. 13.
The microbial rate expression for electron donor utilization, electron acceptor utilization, and
cometabolic.contaminant transformations are added as source and sink terms to the advection-
dispersion. For example, the equation that describes 1-D advective-dispersive transport, the
cometabolic contaminant transformation, and rate-limited sorption results from the addition of
Eq. 8 to Eq. 13 to give:
3C2 n 32C2 8C2 Pb.^p,, px Y.
~^T = Dh~Tl~~v";Tr~7~avKdc2-C2)-Xk2
01 (Jxz OX Q
c2
KS2 + C2 + K?
Equations 12 and 14 must be solved simultaneously, along with similar partial differential
equations for the electron donor (Co), electron acceptor (€A), and the biomass concentration.
These equations and a method for their solutions are given by Semprini and McCarty (1991).
Other solution methods and different model formulations are given in the citations listed in Tables
11 and 12.
This section briefly reviews microbial process models required for modeling cometabolic transfor-
mations. More detailed kinetic models, especially for the cometabolic transformations, are cur-
rently being investigated in the laboratory. Thus, improvements in the models presented here are
likely. Despite these limitations, it is important that models like those presented here be tested to
determine how well they reproduce field observations and to establish the model limitations.
Results of such a modeling exercise will be presented in Section 5, where the results of the Moffett
Field pilot demonstration experiment are presented. In Section 6, results of a preliminary evalua-
tion of in-situ bioremediation at a Superfund site, using the cometabolic transformation will be
presented.
41
-------
SECTION 5
RESULTS OF A PILOT-SCALE STUDY OF ENHANCED
BIOTRANSFORMATION OF HALOGENATED ALKENES BY
METHANOTROPHIC BACTERIA
MOFFETT FIELD STUDY
A pilot-scale field study was performed to assess under field conditions the capacity of native
microorganisms, i.e., bacteria indigenous to the groundwater zone, to degrade halogenated organic
contaminants when proper conditions were provided to enhance bacterial growth. Specifically, the
growth of methanotrophic bacteria was stimulated in a field situation by providing ample supplies
of dissolved methane and oxygen. Under biostimulation conditions, the transformation of repre-
sentative halogenated organic contaminants, including trichloroethene (TCE), cis- and trans-1,2-
dichloroethene (cis- and trans-DCE), and vinyl chloride (VC), was assessed by means of con-
trolled addition, frequent sampling, quantitative analysis, and mass balance comparisons.
This section summarizes the results of the field study. Detailed descriptions of the results are pre-
sented by Roberts et al. (1990) and Semprini et al. (1990,1991).
Field Demonstration Methodology
An effective methodology was developed to evaluate objectively and quantitatively the effective-
ness of the biorestoration approach for stimulating the growth of the desired bacterial population
and transforming the target organic compounds under natural conditions at a field site. The
methodology entails creating a flow field dominated by pumping from an extraction well, while
introducing solutes in known amounts at a nearby injection well and measuring concentrations
regularly at the injection, extraction, and intermediate observation points (Figure 17). Interpreta-
tion of biotransformation behavior could then be made by qualitative examination of the concentra-
tion histories of the various solutes at the several monitoring points, comparing results under bio-
stimulation conditions with results obtained under similar conditions in the absence of biostimula-
tion measures. These interpretations were substantiated by quantitative mass balances.
A specially designed, automated data acquisition and control system (Figure 18), constructed for
this purpose, proved capable of providing continuous records of high-accuracy data over sustained
periods that enabled us to compute mass balances with relative errors of only a few percent.
Details of the system design and operation are presented by Hopkins et al. (1988).
Site Characterization
The site chosen for the field demonstration, at Moffett Naval Air Station (Figure 19), offered a
near-ideal combination of characteristics. The site was representative of a typical situation of
groundwater contamination in the San Francisco Bay area and elsewhere, in which a shallow sand-
and-gravel aquifer is contaminated by chlorinated aliphatic compounds widely used as solvents.
42
-------
I 2
> 4 -
£
I
6 -
SAMPLING
INJECTION WELLS
WELL
EXTRACTION
WELL
SAMPUNG
WELLS
INJECTION
.
CL
1
AY
|/ '//S 1
\ND AND
1
////.
//x///
\S S S S f / / //////////////////////////////////
CLAY
* SI S1 S2 S3 P N3 N2 N1 Nl
i
4 6 8
Distance from well SI, m
10
12
Figure 17. A vertical section of the test zone (from Roberts et al., 1990).
Inject, i on
Pressure
Sw i Lch
Somp1in
Aonifa 1
D i sso1ved
Oxygen
Aeier
Aulii-Channel
Per i sinli i c
Pump
Figure 18. Schematic of the automated Data Acquisition and Control system (from Roberts et al.,
1990).
43
-------
NI
>N2
S3<
S2<
SI
—ms
SCALE, meters
Figure 19. Map of the well field installed at the field site (from Roberts et al., 1990).
Drilling logs revealed that the aquifer at the test site consisted of a layer of silt, sand, and gravel,
approximately 1.2 m thick, at shallow depth (approximately 5 m below the ground surface), well
confined above and below by a silty clay layer of low permeability. The solids exhibited a wide
size range, with approximately 70 wt% > 2 mm and 10 wt% < 0. 1 mm. The organic carbon
content of the aquifer solids was 0.1 1% and the specific surface area was 5 m2/g. Details of
methods used in characterizing the solids are given by Ball et al. (1990).
The formation groundwater was also of appropriate composition for the field experiments. The
water was moderately saline (TDS of 1500 mg/1) and was substantially contaminated by chlori-
nated organic compounds, mainly 1,1,1-trichloroethane, but was devoid of the chlorinated
alkenes-TCE, 1,2-DCE isomers, and VC-chosen as target compounds for this study. There were
no appreciable amounts of toxic metals (Roberts et al., 1989). Nitrate was present in adequate
amounts in the native groundwater (25 mg/1) as a source of nitrogen. Phosphorus concentrations
were low (< 0. 1 mg/1) but near solubility limits of common phosphorus minerals, which probably
were the source of phosphorus.
Sustained pump tests showed that the transmissivity was sufficiently high (approximately
100 m2/day) to permit extracting water at the design rate (approximately 10 1/min) without exces-
sive drawdown at the extraction well. Detailed analysis of the pump tests showed the aquifer
behaved as a leaky aquifer. Model comparisons found a water-table aquitard model best fit the
pump test observations (Johns et al., 1992).
Preliminary tracer tests under natural gradient conditions showed that the local groundwater veloc-
ity was approximately 2 m/day. Preliminary mathematical modeling of the flow field with
RESSQ, (Javandel et al., 1984), imposing a forced gradient on the natural flow field to simulate
injection/extraction operations, showed that injection and extraction rates of approximately 1 1/mi
44
min
-------
and 101/min, respectively, would be sufficient to satisfy the two main requisites for the field
experiment from the hydraulic point of view: 1) complete permeation by injected fluid of the
aquifer in the observation zone between the injection and extraction points (i.e., minimum dilution
by native groundwater in that zone); and 2) complete recovery of the injected fluid at the extraction
well (to assure accurate mass balances).
Extensive bromide tracer tests were undertaken to quantify transport velocities and residence times
in the test zone. Bromide tracer breakthroughs from one of the tests (TR8) are shown in Figure
20. Results summarized in Table 13 confirmed that the aquifer was virtually completely permeated
by the injected fluid in the observation zone, as evidenced by complete steady-state breakthrough
of bromide tracer at the observation wells, under the chosen experimental conditions. Further, the
overall mass balances, comparing the amounts of tracer injected and extracted, demonstrated that
the tracer recovery in the extracted water was essentially complete: after raising the injection and
extraction rates in the second and third seasons of field work, the amount of bromide extracted
agreed within six percent with the amount injected (Table 13). This assured the validity of the
experimental approach, which relied on quantifying the extent of biotransformation of the organic
solutes by comparing instantaneous concentrations at the injection and monitoring points and, over
the long term, by mass balances of the quantities injected and extracted.
The hydraulic residence times (Table 13) between the injection well and the two nearest observa-
tion wells (SI and S2), quantified by tracer tests under the forced gradient conditions, were found
to be in the range of 8 to 23 hr, depending on the pumping rate. The residence time between the
injection and extraction wells was 25 to 40 hr. These residence times were later found to be suit-
able for quantifying the transformation rates of interest in this work. The retardation factors for the
organic solutes, evaluated from relative mobility data obtained in the field, were in the range of two
to ten (Table 14). Details of the results of the tracer tests and modeling the tracer breakthroughs are
given by Roberts et al. (1990) and Chrysikopoulos et al. (1990).
1.2
125
250
375
Time (hours)
Figure 20. Bromide tracer breakthrough and elution in the TracerS experiment, (from Roberts et
al., 1990).
45
-------
TABLE 13. COMPARISON OF BROMIDE TRACER TESTS UNDER INDUCED GRADIENT
CONDITIONS
Test?
Field Season
Injection Rate (1/min)
Extraction Rate (1/min)
Percent Steady-State
Breakthrough
Time to 50% Break-
through (hr)
Percentage Recovered
at the Extraction Well
Well SI
WellS2
Well S3
Extraction
Well SI
WellS2
Well S3
Extraction
TR4
1
1.1
8.0
95
95
80
9
8
20
20
26
66
TR5
1
0.66
8.0
94
72
57
5
9
17
7
20
59
TR8
2
1.36
10.0
100
98
84
13
7.5
16
20
30
105
TR11
3
1.5
10.0
102
100
96
14
9
23
27
40
94
TR12
4
1.5
10.0
100
99
95
15
8
21
26.5
42
ND
°TR4 = Tracer4 experiment, etc.
TABLE 14. RESIDENCE TIMES AND RETARDATION FACTORS FOR THE
CHLORINATED ORGANIC COMPOUNDS BASED ON THE TIME REQUIRED TO
ACHIEVE 50% FRACTIONAL BREAKTHROUGH
Experiment
Tracer4
Tracers
Tracers
Tracerll
Compound
1,1,1-TCA
TCE
TCE
trans-DCE
cis-DCE
TCE
trans-DCE
cis-DCE
Well SI
t50%
(hr)
10
40
60
50
30
50
120
45
Well S2
t50%
(hr)
30
160
150
150
70
175
280
90
R
(SI)
1.3
5
7
6
3
6
13
5
R
(S2)
2.0
9
8
8
4
8
12
4
Tracerl2 Vinyl chloride 13 42 1.6 2.0
46
-------
Sorption
The sorption of the organic solutes by aquifer core samples from the Moffett site was studied in
batch laboratory experiments. Details of these studies and the methods used are given by Roberts
et al. (1989) and Harmon et al. (1990). The studies confirmed that sorption equilibrium was
approximately linear, justifying the use of a distribution coefficient for interpreting and reporting
the sorption equilibrium data. The retardation factors quantified from the field data (Table 14) were
consistent with the results of laboratory studies of sorption. Sorption was strongest for TCE and
weakest for VC, among the compounds studied. The retardation factors calculated from the
laboratory sorption data agreed within a factor of two with those estimated from the transport
experiments conducted in the field. The extent of sorption was approximately equal for all grain
size fractions, but equilibrium was reached much more slowly in large grains than in small ones.
This finding points out that deviations from sorption equilibrium owing to rate limitations may be
an important factor influencing transport behavior. The slow rates of adsorption and desorption
need to be taken into account by incorporating the rate limitation into transport models used for
simulation and design (Section 4).
Results of Biostimulation and Biotransformarion Experiments
The experimental methodology used in the biostimulation and biotransformation experiments is
discussed in detail by Roberts et al. (1990) and Semprini et al. (1990,1991). Biostimulation and
biotransformation experiments were performed under induced-gradient conditions created by the
injection and extraction of groundwater. The well field used for this purpose is shown in Figure
17. The experiments were performed as a series of stimulus-response tests. The stimulus was the
continuous injection of measured concentrations of the chemicals of interest into the test zone, and
the response was the concentration history of the chemicals in the groundwater sampled from the
monitoring wells and the extraction well.
To enhance the effectiveness of biostimulation, methane (primary substrate) and oxygen (electron
acceptor) were dissolved separately in groundwater and introduced as alternating timed pulses.
This was done in order 1) to avoid clogging of the injection well and borehole interface, and 2) to
achieve as uniform a distribution of the microbial growth as possible throughout a substantial por-
tion of the aquifer. The system used for this purpose is shown in Figure 21. Groundwater was
saturated with methane or oxygen using two counter-current gas sorption columns, one for oxygen
and the other for methane. The columns achieved effluent concentrations ranging from 16 to 20
mg/1 methane and 33 to 38 mg/1 oxygen, approximately 80 percent of the saturation values at 20°C,
and atmospheric pressure. The injection solenoids and a pulse tuner permitted the alternated injec-
tion of groundwater containing either methane or oxygen, with varying pulse lengths.
The in-situ biostimulation of a native population of methane-oxidizing bacteria was achieved in
three successive field seasons through the introduction of methane and oxygen dissolved in
groundwater, without any other supplementary nutrients (N and P). Figure 22 shows the concen-
tration history of methane and oxygen at the S2 observation well during the initial biostimulation
experiment. At early time (0 to 100 hr) methane and oxygen behaved like the conservative bromide
tracer, indicating no retardation and minimal consumption. During the period of 200 to 430 hr,
methane and oxygen concentrations rapidly decreased, indicating the population of methane
utilizers had grown to the point of utilizing substantial amounts of methane and oxygen. The ratio
of oxygen consumption to methane consumption was 2.5 g/g, which is significantly lower than
the ratio of 4 that would be required for complete methane oxidation. The lower ratio was also
expected, however, since biological growth assuredly incorporated some of the methane's carbon
into cell mass.
47
-------
Gas
Sorpt i on
Co 1umns
To
Constant
Head
Supp1y
L
O
.
LJ
r
^ $
To
Samp 1 e <-.
AaniFald
/^ ^N
^^3
^
To
I nj ec
Well
1
s
Fi 1 ter Pul se
and Solenoids
Ai xer (^)
D
Disinfection
^ Pressure
Spike Sals Switch
Figure 21. Schematic of the chemical injection system (from Roberts et al, 1990).
100
200 300 400
Time (hours)
500 600
Figure 22. Methane and DO response at the S2 observation well due to the biostimulation of the
test zone (from Semprini et al., 1990).
48
-------
In order to control the clogging of the injection well and borehole interface, the alternate pulse
injection of methane and oxygen containing groundwater was initiated at 430 hr, with a pulse cycle
time of 4 and 8 hr, respectively. The arrival of methane and DO pulses at the S2 well was
observed at a later time. Based on continued methane breakthrough at the observation wells, the
pulsing is believed to have promoted a spatially distributed microbial population in the test zone.
Biofouling of the near well-bore region was thus limited by the pulsing methodology, as antici-
pated in the experimental design; operation run times on the order of six months were feasible,
before clogging forced shutdown and redevelopment of the injection well.
In subsequent field seasons, the uptake of methane and oxygen occurred very rapidly with essen-
tially no lag. The results indicated that some of the methanotrophs stimulated in previous seasons
were present and capable of utilizing methane and oxygen immediately, despite a six month hiatus
since the end of the previous biostimulation. The results indicated that the methanotrophic popula-
tion could survive for long periods without being fed methane, especially under conditions where
no oxygen was present in the aquifer.
In order to evaluate biotransformation, the chlorinated compounds were added to the injected water
(at concentrations in the range of 50 to 100 jig/1), in the absence of methane, until the soil was
saturated, as evidenced by complete breakthrough at the monitoring wells. The feed was then
supplemented with dissolved oxygen and methane. Transformation of the organic target com-
pounds ensued immediately following the beginning of methane utilization, increasing with time as
the bacterial population grew, and ultimately reaching a steady-state value that differed among the
compounds. Figure 23 shows the response of the target compounds at the S2 observation well in
the third season of field testing. The steady-state transformations observed during the third year's
field work (Table 15), quantified by normalization to the bromide fractional breakthrough, were as
follows: TCE, 10 to 19%; cis-DCE, 31 to 47%; trans-DCE, 85 to 90%; and VC, 85 to 95%. Of
100
Time (hours)
150
200
Figure 23. Decreases in normalized concentration of vinyl chloride, trans-DCE, and cis-DCE at
the S2 well in response to biostimulation in the third season (from Semprini et al.,
1990).
49
-------
the values cited, the lower end of the range represents the nearest observation point (1m distant,
8 hr fluid residence time), whereas the upper end of the range represents more distant observation
points with longer residence times (2 to 4 m; 16 to 27 hr). The contaminants' residence times in
the test zone are longer than those of the fluids due to sorption (Table 14). A saturated compound
present as a background contaminant, 1,1,1-trichloroethane (TCA), was not degraded to any
appreciable extent.
TABLE 15. EXTENT OF BIOTRANSFORMATION-THIRD FIELD SEASON
Percent Transformed0
Well VC t-DCE c-DCE TCE
SI
S2
S3
Ext
85
96
95
87
85
90
90
80
31
41
43
47
10
17
19
10
^Estimated by adjusting for bromide-fractional breakthrough.
The field results also indicated that the presence of methane inhibited the rate of transformation of
the chlorinated aliphatics. Periodic changes in methane concentration, which was pulsed, result in
the pulses in concentration VC and t-DCE, which were not pulsed. This effect was most pro-
nounced at the closest observation well SI, where the largest variations in the methane concentra-
tion were observed (Figure 24). High concentrations of VC and t-DCE are associated with high
methane concentrations, indicating slower rates of transformation in the presence of methane.
Transient tests were also performed to determine whether effective transformation rates would be
achieved through the addition of alternative substrates, which unlike methane were not expected to
inhibit transformation rates (Section 3). Figure 25 shows the concentration responses that resulted
when formate and methanol were substituted for methane. Formate and methanol temporally
increased the rates of transformation, while reducing oscillations caused by competitive inhibition.
The ability of formate to keep the system stimulated, thus enhancing transformation rates,
decreased within a few days. Methane was then added to restimulate the population's trans-
formation of the chlorinated aliphatics. When the addition of electron acceptor was terminated, the
concentrations of the chlorinated aliphatic compounds rapidly increased, indicating that the methan-
otrophic population required an input of reducing power (energy) for transformation to be main-
tained. The transient tests demonstrated that the reducing power may be supplied by methane or
other utilizable energy substrates such as formate and methanol.
GC analysis of water samples during active biotransformation of trans-DCE provided evidence of
an intermediate transformation product identified in laboratory studies to be the epoxide of trans-
DCE (Reinhard et al., 1989), which was present in amounts equivalent to a few percent of the
parent compound. The formation of epoxides is a well-known step in the oxidation of aliphatic
compounds (Patel et al. 1982). The epoxides of TCE or VC were not observed, since these
epoxides are much less stable in water and hydrolyze very rapidly, with half-lives on the order of
seconds. The epoxide of trans-DCE hydrolyzes much more slowly in water with a half-life on the
order of 50 to 70 hr (Janssen et al., 1987a and Reinhard et al., 1989) and hence was observed as a
transformation intermediate. No other intermediate products were identified.
50
-------
o
u
100
Time (Hours)
200
Figure 24. Decreases in normalized concentration of vinyl chloride, trans-DCE, and cis-DCE at
the S1 well in response to biostimulation in the third season. Note the pulsing in
concentrations of VC and trans-DCE that result from the pulsing in methane
concentrations (from Semprini et al., 1991).
00
320
520 620
Time (Hours)
720
Figure 25. Response of trans-DCE and cis-DCE at the S1 well to the injection of (1) methane, (2)
formate, (3) methane and formate, (4) methanol, and (5) no electron donor (from
Semprini et al., 1991).
51
-------
In one set of experiments, hydrogen peroxide was applied as a means of increasing the electron
acceptor dose. The addition of hydrogen peroxide permitted operating at a higher rate of methane
feed and increased biological growth but did not enhance the rate of transformation of the target
organic compounds.
Summary
The field experiments at the Moffett site have shown that microbial transformation processes
observed in the laboratory can be promoted and effectively tested in situ, under conditions typical
of many contamination incidents. Stimulation of a specific population of indigenous bacteria that
degrade selected compounds of interest can be accomplished when the proper conditions are
promoted in the subsurface. In this pilot study, the population of methanotrophic bacteria was
enhanced through the addition of methane as a primary substrate for growth.
The biostimulation and biodegradation experiments demonstrated that:
1) A specific class of microorganisms, the methanotrophs, which are indigenous to the subsur-
face environment, can be successfully biostimulated to promote the degradation of certain
chlorinated aliphatic compounds.
2) Partial transformation of VC, 90 to 95%; trans-DCE, 80 to 90%, cis-DCE, 45 to 55%; and
TCE 10 to 20%, occurred over a relatively short flow path of 1 to 2 m in a field test with
fluid residence times of 1 to 2 days.
3) The rate of biotransformation was dependent on the structure of the chlorinated organic com-
pounds, with less chlorinated compounds being transformed more rapidly.
4) An intermediate transformation product, trans-DCE oxide, was produced as a result of trans-
DCE oxidation, which is consistent with the proposed transformation pathway.
5) Methane competitively inhibited the rates of transformation of the chlorinated aliphatics.
6) The substitution of formate and methanol for methane temporally enhanced the rates of trans-
formation of the chlorinated aliphatics.
7) Active utilization of an energy source (methane, formate, or methanol) in the biostimulated
zone was required for chlorinated aliphatic biotransformation to occur.
MODEL INTERPRETATION
The non-steady-state model for simulating the results of the field experiment proved an extremely
useful tool in interpreting the results and comparing them with the laboratory data. A brief descrip-
tion of this model was given in Section 4. More detailed descriptions are given by Semprini and
McCarty (1991,1992). The model incorporated advection, dispersion, sorption with rate limita-
tion, and the microbial processes of substrate utilization, growth, halogenated aliphatic transforma-
tion, and competitive inhibition. The transport was simplified by assuming one-dimensional,
uniform flow, as a computational compromise to permit a more rigorous representation of the
biological processes. Input parameters were estimated based on the results of the laboratory
research of the biological processes, or within ranges of values from the literature, with some
adjustments made to achieve fits to the field results.
52
-------
Figure 26 illustrates a comparison between model simulations and the experimental data from the
field-site's second monitoring well (2.2 m from the injection well) during the first year following
the introduction of both methane and oxygen into the aquifer. The good match obtained indicates
that the basic processes included in the model are good representations of those occurring in the
field. The decreasing concentrations that resulted after about 220 hr (9 days) indicate the growth of
methanotrophic bacteria in the aquifer. An important fitting parameter for the lag phase was the
initial concentration of methanotrophic bacteria, as discussed by Semprini and McCarty (1991).
After about 450 hr, pulsing of methane and oxygen at cycles of 4 and 8 hr respectively, was intro-
duced in order to better distribute the biomass throughout the aquifer.
The simulated increase in methanotrophic biomass computed by the above simulation for a point
2.2 m from the injection well is presented in Figure 27. The biomass rapidly increases during the
period when the methane and oxygen concentrations decrease. A decline in biomass concentration
at approximately 450 hr, due to the lack of methane, is predicted. The decline in biomass concen-
tration slows when the pulsing of methane and oxygen is initiated. The model simulations indi-
cated that the alternate pulsing of methane and DO helped to distribute the biomass in the aquifer
and prevented the bioclogging of injection well and well-bore interface. Additional details of
simulations of biostimulation with pulsing of methane and oxygen are given by Semprini and
McCarty (1991).
When the model was used to simulate the distribution of bacterial mass within the aquifer, the com-
puted results appeared to correspond with the field response, although there was no direct way the
model predictions could be verified. Model simulations of the methanotrophic population response
to restimulation of the test zone in the second and third seasons of testing supported the hypothesis
that the population, once developed, does not decrease in size rapidly, especially under anoxic con-
ditions.
Figure 28 compares model simulations of the degradation of three organics with data from the third
season of testing: a good match to the field observations was achieved. In order to improve the fit
between the model and the field results, the use of both rate-limited sorption/desorption and com-
petitive inhibition kinetics was required. Details of these models are discussed in Section 4. The
rates used in the model simulation for methanotrophic growth, methane utilization, and halogenated
hydrocarbon degradation were in good agreement with those derived under laboratory conditions
that most closely mimicked the field tests (Section 3). One of the main fitting parameters used in
the model calibration was the initial mass of methanotrophs present. Other biological parameters
and rate parameters include stoichiometry between methane and oxygen, Monod rate parameters,
decay coefficient, and values for competitive inhibition. Model parameters are given by Semprini
and McCarty (1991, 1992).
Figure 29 shows the field results and the model simulations of the response of the chlorinated
aliphatics at the nearer observation well (SI) to biostimulation in the third season of field testing.
Model simulations of the effects of competitive inhibition by methane and rate-limited sorption-
desorption also agreed well with the observed dynamic behavior in response to the pulsed injection
of methane and oxygen. Oscillations in t-DCE, and VC concentration resulted from methane
competition, with rate-limited sorption causing greater amplitudes in the oscillations.
The model simulations showed that the responses for the different chlorinated aliphatic compounds
resulted from the compounds having different maximum transformation rates and different sorption
rates and Kj, as summarized in Table 16. The transformation rate parameter values suggest that
vinyl chloride and t-DCE were transformed about as rapidly as methane, whereas c-DCE and TCE
were transformed one and two orders of magnitude less rapidly, respectively. Rates for TCE
53
-------
200
400
600
TIUE (MRS)
Figure 26. Model simulation and observed methane and DO response at the S 2 observation well
during the first season of field testing (adapted from Semprini and McCarty, 1991).
600
TIME (MRS)
Figure 27. Predicted biomass concentration at a node 2.2 m from the injection well due to
stimulation with short and long pulses (from Semprini and McCarty, 1991).
54
-------
60 SO 100 120 140 160 1SO 200
Figure 28. Simulations of the response of methane, VC, t-DCE, and c-DCE, at the S2 well to
biostimulation of the test zone in the third season of field testing. Simulations used
competitive inhibition kinetics and rate-limited sorption (from Semprini and McCarty,
1992).
o
u
z
o
o
2
a:
O
z
40 60 80 100 120 140 160 180 200
Figure 29. Model simulations and the response of methane, VC, t-DCE, and c-DCE at the S1 well
using competitive inhibition kinetics and rate-limited sorption (from Semprini and
McCarty, 1992).
55
-------
TABLE 16. MODEL PARAMETERS FOR SIMULATION OF CHLORINATED ORGANICS IN
BIOSTIM3 SHOWN IN FIGURE 29
Compound
Methane
VC
trans-DCE
cis-DCE
TCE
Kd
(1/mg)
0.0
0.4
1.6
1.6
2.0
a
(d-i)
0.0
0.5
0.3
0.3
0.2
k
(d-1)
2.0
1.5
1.5
0.1
0.02
Ks
(mg/1)
1.0
1.0
1.0
1.0
1.0
k/Ks
(Vmg-d)
2.0
1.5
1.5
0.1
0.02
Kd = sorption distribution coefficient (1/mg).
a = rate coefficient for sorption (d*1).
k = maximum transformation rate (d'1).
Ks = half-saturation coefficient (mg/1).
transformation (k/Ks) are in the range of those determined in the laboratory under non-optimal
growth conditions (Table 9).
The model simulations of the Moffett Field studies demonstrate the usefulness of models that
incorporate fundamental microbial and transport processes. The processes of microbial growth,
utilization of methane and oxygen, competitive inhibition of the cometabolic transformation, and
rate-limited sorption, must be considered along with advection and dispersion. In these simula-
tions, the model was greatly simplified by assuming 1-D transport, as well as a shallow biofilm;
even with these simplifications, the model simulated most aspects of the field test quite well.
However, 2-D and 3-D models may be necessary at other locations, where conditions are more
complex, assuming this can be justified by the availability of sufficient data on aquifer character-
istics. The matter of adequately characterizing highly heterogeneous sites for purposes of bio-
remediation assessment is so challenging as to be virtually intractable, given the limits of present
knowledge. In any case, mathematical model representations of such highly complex situations is
beyond the scope of this paper.
This model that was developed, validated, and calibrated in the Moffett Naval Air Station field
study will be used in preliminary evaluations of bioremediation scenarios for a Superfund site in
Michigan. These model simulations are presented in Section 6.
56
-------
SECTION 6
FEASIBILITY STUDIES FOR A SITE
INTRODUCTION AND OBJECTIVES
The Moffett field evaluation demonstrated that indigenous microorganisms could be stimulated
through injection of methane and oxygen to degrade TCE, DCE, and VC. As part of that evalua-
tion, methodologies for determining the potential for methanotrophic in-situ biorestoration of sites
contaminated with chlorinated solvents were developed. In order to illustrate the use of these
methodologies, to further evaluate the potential of this process for bioremediation of contaminated
groundwater, and to better consider the engineering factors involved in system design, a contami-
nated site was sought that would provide a full-scale test case for the process. The St. Joseph
Superfund site appeared to be a good candidate for this evaluation.
At the St. Joseph site, the aquifer is relatively homogeneous and heavily contaminated with the
same three contaminants for which degradation by a methanotrophic community has been demon-
strated at the Moffett site. Trichloroethylene (TCE), 1,2-cis-dichloroethylene (c-DCE), 1,2-trans-
dichloroethylene (t-DCE), and vinyl chloride (VC) are present at groundwater concentrations from
0.8 to 8.0 mg/1 (Figure 30). It appears from the data available that the latter compounds were
formed through anaerobic transformation of TCE as described under Section 3 on Biotransforma-
tion. These compounds were also found at the Moffett test site to be more readily degraded than
TCE.
The specific objectives of this evaluation and the methodologies developed were to:
• Determine whether a native population of methanotrophic bacteria existed in the St.
Joseph aquifer, and if so, to evaluate this population's ability to degrade the contami-
nants of concern,
• Determine the degree of sorption of the contaminants to the aquifer material.
• Evaluate likely designs for in-situ treatment through use of the non-steady-state model
developed for this process.
To simulate results for in-situ bioremediation, the presence of an indigenous population and its size
must be known. This is the basis for the first objective. Knowledge of the sorptive properties of
contaminants at a given site is also important for any bioremediation scheme as this affects how the
contaminants move relative to the water. The rates of biodegradation are also functions of the
amount of compound sorbed, as well as that in solution.
The third task was to evaluate possible designs for the remediation system. The model developed
and verified for simulation at the Moffett test site (Section 5) was believed suitable for initial
evaluations at the St. Joseph site. In addition, a regional flow model was developed to evaluate the
important parameters affecting flow of water into and through the St. Joseph aquifer system.
57
-------
Figure 30. Distribution of TCE, DCE, and VC at the St. Joseph site (after Keck, 1988).
PROCEDURES
Collection of Aquifer Samples
The U.S. Environmental Protection Agency's (EPA) Robert S. Kerr Environmental Research
Laboratory provided their drilling crew and drilling rig for collection of aquifer cores according to a
protocol designed to avoid sample contamination (Wilson and Leach, 1989). Samples were taken
near sampling wells OW-18 and OW-29 (Figure 31), where the maximum concentrations of DCE
and VC were found in the western portion of the plume, migrating toward Lake Michigan. Here,
the groundwater lies about 10m below the surface, and the center of the contaminant plume is
about 20 m below the surface.
58
-------
St. Joseph Site
Lake Michigan - 17«J m »i>.
Figure 31. Vertical profile of subsurface at St. Joseph (from McCarty et al., 1991).
Aseptically obtained samples of aquifer material were collected using a 15-cm diameter hollow-
stem auger equipped with a cover plate to exclude the entrance of aquifer material into the hollow
stem during drilling. When the auger reached the desired level, a 10-cm pitcher barrel, equipped
with a plunger to prevent loss of aquifer material, was driven first to open the cover plate and then
into previously undisturbed aquifer material below the auger (Wilson and Leach, 1989). The
pitcher barrel was then removed from the auger, plugged with a rubber stopper to prevent loss of
material, and attached by the upper end in a horizontal position to a drive device. The stoppered
end of the barrel was placed in a glove box, nitrogen from a cylinder was allowed to fill the glove
box to exclude oxygen. The stopper was removed, the drive device was activated to force the
plunger in the barrel to move forward a short distance, exposing about 10 cm of the core which
was tapped free, leaving a surface of fresh aquifer material. A sterile circular cover plate with a
hole in the center was then attached to the barrel. The auger plate had a sharp circular edge around
the hole that permitted only the central material in the core to pass, while excluding the exterior core
material that had come into contact with the barrel wall. The central portion of the core, which was
free from extraneous contamination due to sampling, was forced out by activating the drive device
and was collected in sterile 1-liter Mason® jars for transport to the Stanford laboratory, where they
were stored in the dark at 4°C until used. Because of difficulties with the fine sandy aquifer
material, aseptically collected core samples were obtained only from the upper 4 m at OW-29 and
the upper 2 m of the aquifer at OW-18 in initial sampling. Later, the cover plate on the auger was
modified to prevent the intrusion of fine sand during drilling; and samples from near the center of
the plume were obtained. Material was also obtained between 10 and 12 m below the water
surface near OW-29 for physical characterization.
Column Microcosm Studies
The objectives of column microcosm studies were 1) to determine whether indigenous methano-
trophic bacteria are present at the St. Joseph site, 2) to quantify the time required to increase the
population of such bacteria to an adequate level for bioremediation, and 3) to assess the capability
of such a stimulated population to degrade TCE, t-DCE, and VC.
59
-------
The 2-cm diameter glass column microcosms (Figure 32) had a volume of 133 ml, contained about
200 g of aquifer material, had a void volume of 50 ml, and were constructed and operated as
described previously (Roberts et al., 1989, Lanzarone and McCarty, 1990). The microcosms were
autoclaved and aquifer material was aseptically added, as was filter-sterilized groundwater from an
uncontaminated location at the St. Joseph site. The columns were operated by exchange of column
fluid after periods ranging from a few to 30 days. Here, replacement of column fluid was effected
by upward flow of about 140 ml of amended and filter-sterilized groundwater over about a 30-min
period (Figure 33).
Five columns (B, C, D, E, F) contained aquifer material from near well OW-18 and two (G, K)
from near well OW-29. Column flow characteristics were determined by exchange of water
containing a bromide tracer, the concentration of which was monitored by ion chromatography
(Figure 34). Then, some columns were exchanged with water containing only methane and
oxygen to determine whether methane-consuming methanotrophs were present. Once stimulation
of methanotrophs was in evidence, VC was added to the exchange water and the possible degrada-
tion of VC was determined. Later, TCE, t-DCE, as well as VC were added to evaluate degradation
potential.
0
47 cm
Teflon Connecton
and End Plugs
Teflon Tubing
Rubber Stopper
Steel Needle
Water
Glait Wool
Aquifer Material
Total Volume
Vt a 133 ml
Pore Volume
Vw« 50 ml
Glaas Column
1.0. = 2 cm
Rubber Stopper
Figure 32. Schematic diagram of column microcosms (from Lanzarone and McCarty, 1990).
60
-------
Influent Sample
Syringe Pump
5 ml/min
Two 100 ml
Syringes
n
0-1B ml
Initial Effluent
19-106 ml
Discarded
106-126 ml
Final Effluent
Figure 33. Schematic figure of procedure used in exchanging column fluids (from Lanzarone and
McCarty, 1990).
1.2
i.o-
o 0.8-
u
u
M
a
i
0.6-
S 0.4-
0.2
0.0*
**«.
• e
I
-0
20 40 60 80 100
Uolume Pumped, ml
9 Col A • ColH
a ColB A Coll
• ColC x ColJ
A ColD • ColK
* CdE • ColL
• ColF M ColM
• Col Q • Col N
120
Figure 34. Breakthrough curves of bromide tracer for the different columns.
61
-------
COLUMN MICROCOSM RESULTS
Bromide Tracer
No breakthrough of the nonreactive and nonsorptive bromide tracer was observed in preliminary
column microcosm studies until the passage of about 30 ml of fluid. All effluents reached the
influent concentration after about 90 ml passed through (Figure 34). The estimated average
column pore volume is about 50 ml. The sharp bromide breakthroughs for all columns are consis-
tent with the uniform and fine grain size of the aquifer material. These results indicated that the
first 30 ml of column effluent was representative of concentrations present within the column and
the absence of the influence of short-circuiting by the feed solution. Thus, the first 30 ml from
each column during each exchange were taken for analysis to represent column fluid characteristics
before the exchange. The results also indicated that a total exchange of 90 ml of fluid during each
period was required to insure that the influent materials were distributed throughout the columns.
Methane and Oxygen Consumption
Methane (about 3.5 mg/1) and oxygen (about 25 mg/1) were added initially to four columns (C, E,
G, K), which were then exchanged similarly at various times to determine whether methane con-
sumption had begun. Methane consumption in Column C was complete within 31 days, and in
Column E within 41 days (Figure 35). The columns contained aquifer material taken near well
OW-18, but at somewhat different depths. No data reveal methane utilization before day 28, but
methane was consumed in Column C by day 31. Figure 35 illustrates the averages of concentra-
tions in two samples taken at each time period, one for the first 15 ml and the other for the second
15ml.
The time intervals for the onset of methane consumption in Columns G and K, with aquifer
material from near well OW-29, were similar, although initially methane consumption was not
uniform throughout the columns. Methane concentration in the initial effluent was little changed
from the influent values; but in the sample from the second 15 ml, it was nearly completely
removed in both columns. This suggests that the methanotrophic population may not have been
evenly distributed throughout the columns. Since the activity was higher in both columns at the
10 20 30
Tim*, day*
«o so
Figure 35. Methane removal in column microcosms (from McCarty et al., 1991).
62
-------
influent end, it is also possible that the columns might have become contaminated with methano-
trophs during an exchange. Assuming this did not occur, the population levels in all the columns
appeared similar. The methanotrophic populations in St. Joseph aquifer materials were much
lower than in Moffett material, which exhibited noticeable methane consumption within two
weeks. Since the doubling time for methanotrophs is about one day, this difference indicates that
the St. Joseph methanotrophic population is four to five orders of magnitude smaller than the
Moffett population. The previously mentioned observation that the distribution of methanotrophic
bacteria appears patchy would be consistent with this finding that the indigenous population of
methanotrophs is extremely sparse.
Vinyl Chloride Degradation
After methane consumption occurred, approximately 0.175 mg/1 vinyl chloride (VC) together with
methane and oxygen were added to Columns C and E through fluid exchange. The control
Column B was then exchanged with fluid containing dissolved oxygen and VC, but no methane.
Periodically, the pore fluid was exchanged with similarly amended fluid, and the initial effluents
were analyzed for dissolved oxygen, methane, and VC. The VC results are presented in Figure
36. Time zero represents the time of initial VC addition. At the first exchange one week after VC
addition, 20 percent of the VC in the control column was removed, probably from sorption to the
aquifer material. In the methane-stimulated columns, however, 75 percent VC removal was
obtained. The second exchange, carried out one week later, showed no VC removal in the control
column, indicating that VC had reached sorption equilibrium. Columns C and E both showed
increased VC removal to about 80 percent, possibly as a result of the increasing size of the methan-
otrophic population. Complete methane consumption was obtained between each exchange in all
cases.
The influent VC concentration for the third exchange (t = 28 days) was increased to 0.78 mg/1 to
determine whether higher VC concentrations similar to those found at the St. Joseph site would be
inhibitory. The results again showed about 80 percent VC removal in Columns C and E over a
period of two weeks and only small sorptive removal in the control column. The columns were
exchanged once a week for two additional weeks while maintaining a higher influent concentration.
The exchange results (t = 35 and 42 days) showed a consistent 75 to 80 percent VC removal from
Columns C and E with little or no removal in the control column. These data indicate that the
higher VC concentrations were not toxic, and that the degree of removal with higher concentrations
was similar to that with lower concentrations.
Following the above exchanges (i.e., t > 42 days), the influent VC concentration was reduced to
0.28 mg/1 and left in the columns for two weeks. After this period, the control column showed VC
concentrations 20 percent greater than the influent concentration resulting from desorption of VC
from the aquifer solids equilibrated previously at the higher VC levels. Both Columns C and E
showed 95 percent VC removal during this period. When analyzed 30 days later, VC was
removed to below the detection limit in Columns C and E. However in control Column B, it was
two times greater than the influent concentration because of the desorption of VC from the aquifer
solids, which had been equilibrated previously at higher VC levels. This confirms that methane
consumption was required for VC degradation to occur.
Trans-1.2-DichloroethyIene and Trichloroethylene
On day 54,0.16 mg/1 t-DCE and 0.07 mg/1 TCE were added together with 0.1 mg/l VC to the feed
for Columns B, C, and E. The results are presented in Figure 36, with time zero for the lower two
graphs representing day 54 on the upper graph when t-DCE and TCE were added. A comparison
63
-------
o.s
(A)
(B)
o
0.2
0.30
0.25
-I
0> 0.20-
Ul
o.is
0.10
0.05
0
Influent
Control
ColEEfll
ColCEfll
10 20 30 40
Tlm«, days
30 60
(C)
0.14
0.12
0.10
0.01
0.04
0.04
• .02
0
Control
ColEEfll
Col C Em
10 20 30 ' tO X 60
Tlm», days .
Figure 36. VC, t-DCE, and TCE removal in column microcosms. The origin on the time scale in
(B) and (C) corresponds to t = 54 days in (A) (from McCarty et al., 1991).
between the t-DCE concentration in the effluent from the control column and Columns C and E
suggests a t-DCE removal of 60 percent by the methanotrophs. TCE levels were the same in
effluents from Columns C, E, and the control.
64
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Influent concentrations of t-DCE and TCE were then increased to 0.25 and 0.1 mg/1, respectively,
and left in the columns for one week. Exchange results showed removals of 60 to 75 percent for
VC, 35 to 40 percent for t-DCE, and 0 percent for TCE in the active columns. The columns were
fed again and left for 21 additional days, after which the active columns showed removals of 95
percent for VC, 80 percent for t-DCE, and 30 percent for TCE. The extent of VC transformation
before and after the addition of t-DCE and TCE to the columns appears to be approximately the
same.
Long-Term Vinyl Chloride Results
One study was conducted to simulate field conditions in which methane, oxygen, and VC would
be mixed together initially and allowed to incubate in the aquifer. Here, VC was added to two
previously unexchanged columns and left to incubate at room temperature. Dissolved oxygen (25
mg/1) and methane (4.3 mg/1) were added with VC (0.95 mg/1) to Column D, but only oxygen and
VC were added to control Column F. After ten weeks of incubation, the columns were exchanged
and the initial effluents analyzed. The control column showed a 17 percent decrease in VC, prob-
ably from sorption, whereas the methanated column showed 90 percent removal of VC. All the
methane and about 15 mg/1 of the dissolved oxygen were consumed in the methanated column.
The theoretical oxygen demand from the methane and VC is in the range of 13 to 18 mg/1 (depend-
ing on the oxygen-to-methane ratio assumed), which brackets the actual oxygen usage. In the
control column, about 6 mg/1 dissolved oxygen was used over the ten-week period. This test
illustrates the potential effectiveness of a single aquifer exchange for removing a major portion of
the VC contamination in place. VC removals in the range of 95 percent were obtained when three
weeks rather than one week was allowed between exchanges. The fact that additional removal can
be obtained after the depletion of methane is consistent with other column studies conducted at
Stanford (Roberts et al., 1989).
REMEDIATION SCENARIOS
Model simulations will be presented to show how biotransformation models can be useful tools in
preliminary designs of in-situ and on-site bioremediation processes. As a case study, model simu-
lations were performed to investigate biorestoration schemes for the St. Joseph site using the
numerical code that had been developed, validated, and calibrated in the Moffett Naval Air Station
field study (Sections 4 and 5). The model includes the transport processes of advection, disper-
sion, and sorption combined with microbial processes for the growth of methanotrophic bacteria,
utilization of methane as the growth substrate and oxygen as the electron acceptor, and microbial
kinetics for the cometabolic transformation of chlorinated aliphatic compounds.
The cometabolic transformation of chlorinated aliphatic compounds by methanotrophs is a fairly
complex process with many factors contributing to the effectiveness of this form of treatment.
Such cometabolic transformation is governed by competitive inhibition kinetics, where high
methane concentrations can reduce die transformation rates of the cometabolized substrates. The
biotransformations are also limited in the absence of oxygen, thus the model reaction kinetics also
include an expression for oxygen concentration. The ability of methanotrophs to degrade chlori-
nated aliphatic compounds is closely associated with their utilization of methane. During periods
when methane is not being consumed, the ability to transform chlorinated compounds is rapidly
lost. The model also incorporates this activity-loss process through a first-order deactivation rate
expression.
The simulations used input parameters obtained from the calibration with the Moffett Field test
results (Roberts et al., 1989, Semprini and McCarty, 1991, Semprini and McCarty, 1992).
Additional parameters obtained from the St. Joseph study were also entered into the simulation
65
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model. These parameters include the extent of sorption onto the aquifer solids and the estimate of
initial concentration of methanotrophs in the treatment zone. The latter was estimated through the
model by selecting a population that would best provide a simulation of the column microcosm
results. Simulations were then performed using rate coefficients derived from the Moffett Field
study to determine whether the Moffett-derived parameters were consistent with results from the
laboratory column studies with St. Joseph aquifer material. This served to help verify the applica-
bility of the Moffett results to the St. Joseph site.
The model described above has been shown to be capable of reproducing all aspects of the Moffett
Field experiments (Section 5). The model, which was calibrated using the results from the Moffett
study, thus provided a strong framework for conducting simulations of in-situ treatment for the St.
Joseph site. Incorporation of the specific measured sorption parameters (K& a) for the chlorinated
aliphatic compounds of interest on St. Joseph aquifer solids, die initial methanotrophic population
appropriate for the St. Joseph site, and transformation rate coefficients consistent with the
laboratory studies, provides the necessary information for the model suitable for the St. Joseph
site evaluations.
This preliminary modeling study did not attempt to devise optimal designs for the in-situ remedia-
tion system. Remediation criteria and adequate information on site hydrogeology that are needed
for developing an optimal design were not available. Rather, the process design developed here is
one of perhaps several feasible configurations that might be used for comparison with alternative
clean-up strategies, such as simple pump-and-treat remediation.
The remediation system for which model simulations of bioremediation were performed (Figure
37) is conceived as a pump-and-treat system. However, ah1 of the treatment is carried out biologi-
cally in the aquifer itself. Here, groundwater is extracted through a series of wells spanning the
contaminant plume in a direction perpendicular to that of groundwater flow. At the surface,
methane and oxygen are added to the extracted groundwater, either together or in alternating
pulses, depending upon the extent to which a stimulated methanotrophic population has been
developed. The alternating pulses are used to distribute the microbial growth throughout the test
zone. The groundwater containing the methane, oxygen, and extracted contaminants is reinjected
into the treatment zone through a series of downgradient monitoring wells, distributed parallel to
the extraction wells. In the subsurface biotreatment zone, both the in-place contaminants and the
reinjected contaminants are biologically degraded.
The above in-situ biotreatment system was directly compared with a pump-and-treat system as
illustrated in Figure 38. In the latter case, an unspecified form of surface treatment is assumed to
remove contaminants quantitatively before the water is reinjected. Otherwise, the two systems are
operated identically, and thus can be compared directly through model simulations.
For the two alternate simulations, a complete cross-section segment through the contaminant plume
was chosen for treatment. The cross section assumed was 200 m wide, which is expected to span
the width of the plume. The length and depth of the assumed remediation segment are 120 m and
20 m, respectively. In this segment the contaminants VC, c-DCE, and t-DCE were assumed each
to be initially present at a uniform aqueous concentration of 1 mg/1, which is within the general
range of observed concentrations.
The simulations were performed using rate parameters from the Moffett study (18°C) and the labo-
ratory study with St. Joseph material (22°C), that were adjusted for the lower temperature of 10°C
for the St. Joseph aquifer. Based on the laboratory sorption study and the comparison with
66
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CHLORINATED
ETHENES REINJECTED
INTO THE
BIOSTIMULATED
ZONE
GROUNDWATER
FLOW
BIOSTIMULATED
ZONE
METHANE AND OXYGEN
ADDITION IN ALTERNATING
PULStS
Figure 37. In-situ bioremediation case simulated (from McCarty et al., 1991).
ABOVE GROUND TREATMENT
CLEAN
GROUNDWATER
REINJECTED
GROUNDWATER
FLOW
EXTRACTION
Figure 38. Pump-and-treat system for comparison with bioremediation (from McCarty et al.,
1991).
67
-------
Moffett Field results, a retardation factor for VC of 1.7 was assumed. Based upon results sum-
marized in Table 15, c- and t-DCE were assumed to be more strongly sorbed, with retardation fac-
tors for both of 4.
The injection and extraction flow rates used in the simulation model provide a fluid residence time
in the treatment zone of 50 days. This residence time was estimated based on the laboratory
column estimates of initial methanotrophic population and a time of 30 to 50 days for significant
growth to occur. At the colder St. Joseph groundwater temperature, this time would probably be
twice as great. Thus, the chosen residence time of 50 days should be sufficient to permit the mixed
methane and oxygen in the injection water to completely permeate the treatment zone before
observable methane and oxygen uptake occurs. After the breakthrough of methane and oxygen
occurs at the extraction well(s), the induced flow (pumping) is stopped to wait for the onset of
methanotrophic growth and the initial transformation of the chlorinated compounds to occur. After
that, injection would again commence, but with alternating pulses of water containing methane and
oxygen.
Figure 39 shows the results of the model simulation for VC treatment. Shown are the concen-
trations of VC and methane as a function of time at a node adjacent to the extraction well. The
treatment is operated initially as a batch process as described above. Groundwater is extracted and
methane and oxygen are added to the reinjected fluid until breakthrough of methane and oxygen is
observed at the extraction well after about 40 to 50 days. To minimize both the amount of water
extracted and injected and the quantities of methane and oxygen added, all pumping is then
stopped. The growth of methanotrophic bacteria is enhanced through consumption of the injected
methane and oxygen, and at the same time, the cometabolic degradation of the VC begins. After
100 days (50 days of no pumping), all of the methane should be consumed; and the model indi-
cates that 80% of the VC will be biodegraded. The simulations show that despite the presence of
methanotrophic bacteria, once methane is completely utilized, the transformation of VC rapidly
ceases. The laboratory column-study results, however, suggest that VC transformation may
continue at a lesser rate, so that as much as 90% or more may actually be decomposed. Thus the
model simulations may somewhat underestimate the actual transformations. Nevertheless, in line
with the model simulations, once methane is depleted in the aquifer, more methane and oxygen are
then added to the treatment zone to accomplish additional degradation. Here, methane break-
through occurs once again after 120 to 140 hr of injection, and the subsequent biodegradation
brings VC to the drinking-water standard.
The results of this simulated, semi-continuous batch treatment of the test zone agree well with the
results of the batch column studies presented earlier. In the column studies the initial batch
addition of 3.5 mg/1 of methane resulted in approximately a 75% reduction in VC concentration. A
slightly greater reduction was calculated in the field treatment simulation. However, in the column
studies with longer exchange periods, VC decomposition of 90% occurred. Thus, the column
results bracket the simulations. The combined results indicate that the rate coefficients derived
from the Moffett Field study, when used in the model simulations, agree well with the results of
the laboratory experiments with St. Joseph materials. This indicates that the model simulations
were performed with a reasonable set of input parameters for the St. Joseph site.
The above bioremediation simulation system is compared with a simulation of a conventional
pump-and-treat approach in Figure 40. Here, the systems were operated exactly the same with
respect to hydraulic conditions, that is, with the same amount of pumping and over exactly the
same time periods. The biostimulation treatment is shown to be slightly better than the pump-and-
treat method, based on the pumping time required to reduce VC concentration in the treatment
zone. Since VC is weakly sorbed (R = 1.7), a pump-and-treat system should also work fairly
well. The effect of possible aquifer heterogeneities are not included in the simulations, however.
68
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(EXTRACTION NODE)
I
z
o
u
I
o
o
20
180 200
Figure 39. Simulation response to biostimulation with methane for VC remediation.
(EXTRACTION NODE)
20 40 60 80 100 120 140 160 180 200
Figure 40. Comparison of in-situ bioremediation of VC with pump-and-treat. Biostim+pump is a
combination of surface treatment and in-situ treatment (from McCarty et al., 1991).
69
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Heterogeneities could result in significant tailing in the pump-and-treat response that might make
the bioremediation method even more effective by comparison. Heterogeneities may also affect the
efficiency with which 02, CFU, etc., can be distributed within the aquifer and reduce the effective-
ness of bioremediation as well.
The model simulations illustrated in Figure 40 also indicate the advisability of recycling the con-
taminants through the treatment zone rather than removing them through treatment at the surface.
The dashed line with triangles (Biostim+pump) shows a simulation in which the contaminants in
the extracted water are removed using surface treatment before reinjection. Methane and oxygen
are added to the surface-treated reinjected water. Some, but not a significant, enhancement of
removal is achieved by adding surface treatment. Moreover, the in-situ bioremediation process
also degrades the reinjected contaminants to nontoxic end products, which is an advantage over
some forms of surface treatment.
Simulations for t-DCE are shown in Figure 41. In the Moffett Field study, t-DCE had the same
degradation rate coefficient as VC, but was more strongly sorbed onto the aquifer solids. This
resulted in a slower rate of treatment for t-DCE. The Moffett transformation rate coefficient
(corrected for temperature) was used in the St. Joseph simulations, along with a sorption retarda-
tion factor of 4.0. The simulation model indicates that a 50% reduction in t-DCE concentration
would occur during the initial methane injection phase. As in the Moffett study, the lower reduc-
tion with t-DCE than with VC results because of stronger sorption by the former compound. The
simulation results, however, suggest that in-situ bioremediation for t-DCE is more attractive than
pump-and-treat alone. Thus, some degree of sorption to keep compounds in the aquifer favors in-
situ bioremediation.
The corresponding simulation for c-DCE is illustrated in Figure 42. In the Moffett Field study
c-DCE was degraded at a rate an order of magnitude lower than t-DCE. With this lower rate, a
longer in-situ treatment period is required to pump and treat. However, the c-DCE that is biode-
graded within the aquifer is permanently destroyed, and thus in-situ treatment has important bene-
fits. Also, the proposed drinking-water MCL for c-DCE is much higher than for VC. Thus, treat-
ment that is adequate for bringing VC and t-DCE into compliance may be adequate for c-DCE as
well. This possibility will not be known until clean-up requirements are established. Nonetheless,
the simulation does demonstrate that in the treatment of mixtures of compounds by in-situ
bioremediation, reaction rates and degrees of sorption, as well as remediation goals, must all be
considered before the best strategy can be formulated.
A summary of the engineering aspects of the in-situ bioremediation scheme, as described above
and illustrated in Figure 37, and based upon the assumptions already described is contained in
Table 17. For this case, a total of 480,000 m3 of aquifer would be treated in 400 days, using a
total extraction rate of 400 gpm. A total of 1375 kg of the original 1617 kg of contaminants pres-
ent in the treated volume would be biodegraded to nontoxic end products. This treatment would
require 5,200 kg of methane and 19,200 kg of oxygen.
These simulations show how model simulations provide a support for evaluating in-situ processes
at a given site. Simulations permit comparisons between remediation via bioremediation, biore-
mediation in combination with pump-and-treat, and pump-and-treat. For the example shown,
pump-and-treat might be satisfactory for remediation as well. Here, a suitable treatment process at
the surface must be used. The hydrogeology was also greatly simplified, and effects of aquifer
heterogeneities on pump-and-treat and bioremediation were neglected. In order to make the best
selection, however, more information is required on the hydraulic and transport characteristics of
the site, methanotrophic population distribution, and remediation goals, as well as other site-related
clean-up criteria.
70
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I
I
I
I
o
Figure 41. Comparison of in-situ bioremediation of t-DCE with pump-and-treat.
X
9
800
Figure 42. Comparison of in-situ bioremediation of c-DCE with pump-and-treat (from McCarty et
al., 1991).
71
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Initial Initial
Aqueous (kg) Solids (kg)
VC 158 117
t-DCE 158 513
c-DCE 158 513
Methane and Oxvgen Requirements
Initial
Total (kg)
275
671
671
TABLE 17. ENGINEERING SUMMARY OF IN-SITU BIOLOGICAL TREATMENT
(from McCarty et al.. 1991)
Dimensions of the Treated Zone
120mx200mx20m = 480,000 m3
length width depth volume
Mass Biodegraded in 400 Days (1 mg/1 initial aaueous)
Treated Percent
(kg) Treated
274 > 99
665 > 99
436 65
Methane 5,200 kg
Oxygen 19,200 kg
Extraction and Injection Rates
1.8m3/min(400gpm)
It should be emphasized again that no attempt was made to develop an optimized treatment system
with the simulations since insufficient information is presently available to do so. The simulations
serve primarily to indicate the factors involved in an in-situ treatment scheme and the time scales
that must be considered. Since in-situ bioremediation in this way has not been attempted before,
there are many possibilities for improvement on the design assumed for the simulations. The
simulations do help to bracket the probable range of outcomes, and thus should help enormously in
the design and interpretation of demonstrations, such as the one planned for the St. Joseph site.
72
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
REPORT NO.
EPA/600/R-92/042
3. RECIPIENT'S ACCESSION NO
PB92-146 943
TITLE AND SUBTITLE
METHODOLOGIES FOR EVALUATING IN-SITU BIOREMEDIATION OF
CHLORINATED SOLVENTS
5 REPORT DATE
March 1992
6. PERFORMING ORGANIZATION CODE
AUTHOR(S)
L. SEMPRINI, D. GRBIC-GALIC, P. McCARTY, & P. ROBERTS
8. PERFORMING ORGANIZATION REPORT NO.
PERFORMING ORGANIZATION NAME AND ADDRESS
DEPARTMENT OF CIVIL ENGINEERING
STANFORD UNIVERSITY
STANFORD, CALIFORNIA 94305
10. PROGRAM ELEMENT NO.
TEKY1A
11 CONTRACT/GRANT NO.
CR-815816
2. SPONSORING AGENCY NAME AND ADDRESS
ROBERT S. KERR ENVIRONMENTAL RESEARCH LABORATORY
U.S. ENVIRONMENTAL PROTECTION AGENCY
P.O. BOX 1198
ADA, OK 74820
13. TYPE OF REPORT AND PERIOD COVERED
RESEARCH REPORT 8/21/89-6/14/91
14. SPONSORING AGENCY CODE
EPA/600/15
5. SUPPLEMENTARY NOTES
PROJECT OFFICER: STEPHEN G. SCHMELLING FTS: 743-2,434
6. ABSTRACT
This report summarizes the behavior of and requisite conditions for a class of natural biological processes that
can transform chlorinated aliphatic compounds. These compounds are among the most prevalent hazardous
chemical contaminants found in municipal and industrial wastewaters, landfills and landfill leachates, industrial
disposal sites, and groundwater. Biological degradation is one approach that has the potential for destroying
hazardous chemicals so that they can be rendered harmless for all time. Methodologies are presented that are
useful for evaluating the potential for biorestoration of groundwater contaminated with chlorinated aliphatic
compounds. The report is composed of six sections. Section 1 provides an introduction and an overview of the
problems with chlorinated aliphatic compounds in groundwater. Section 2 presents a review of the processes
affecting the movement and fate of chlorinated aliphatics in the subsurface, including advection, dispersion, sorption
and relative mobility, diffusional transport, and immiscible transport. Section 3 provides a thorough review of the
microbial transformation of organic pollutants. Basic microbial metabolic processes are reviewed, focusing on an
aerobic and aerobic transformations of chlorinated aliphatic compounds. Laboratory studies of aerobic cometabolic
transformation and degradation of TCE by methanotrophs and methanotrophic communities are summarized. In
Section 4 transport and microbial process models are presented and incorporated into a model for the aerobic
cometabolic transformation of chlorinated aliphatics by methanotrophic communities. Section 5 presents pilot-scale
results of enhanced in-situ biotransformation of halogenated alkenes, including TCE, cis- and trans-DCE, and vinyl
chloride by methantrophic bacteria along with model simulations of the results. Section 6 presents an example
study to evaluate the potential and limitations for groundwater bioremediation at a Superfund site by
methanotrophs. Methodologies and results are presented for evaluating the presence of a native methantrophic
community and its ability to degrade the contaminants of concern; determining the sorption of contaminants to the
aquifer material; and preliminary designing of an in-situ treatment approach using the model previously described.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
S.IDENTIFIERS/OPEN ENDED TERMS
IN-SITU BIORE1EDIATION*
IN-SITU PROCESSES
CHLORINATED SOLVENTS
AEROBIC TREATMENT
BIOLOGICAL DEGRADATION
BIORESTORATION
*Sugeest Addition to List
TCE
DCE
VINYL CHLORIDE
GROUNDWATER
BIODEGRADATION
BIORESTORATION
BIOTRANSFORMATION
COSAT! field,Group
18. DISTRIBUTION STATEMENT
RELEASE TO THE PUBLIC
EPA Form 2220-1 (R«v. 4-77) PREVIOUS EDITION is OBSOLETE
19 SECURITY CLASS i fins Report)
UNCLASSIFIED
20. SECURITY CLASS (This pu
UNCLASSIFIED
96
22 PRICE
86
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