HYDROLOGICAL PROCESSES
Hydrol. Process. 13, 2847±2859 (1999)

Numerical simulation of transport and sequential
biodegradation of chlorinated aliphatichydrocarbons using CHAIN_2D
J. Schaerlaekens,1* D. Mallants,2 J. SÆimuÃnek,3 M. Th. van Genuchten3 and J. Feyen1
1Laboratory for Soil and Water, K.U. Leuven, Leuven, Belgium
2Belgian Nuclear Research Centre SCK-CEN, Mol, Belgium
3

USDA Soil Salinity Laboratory, Riverside, CA, USA

Abstract:
Microbiological degradation of perchloroethylene (PCE) under anaerobic conditions follows a series of chain
reactions, in which, sequentially, trichloroethylene (TCE), cis-dichloroethylene (c-DCE), vinylchloride (VC)and ethene are generated. First-order degradation rate constants, partitioning coecients and mass exchange
rates for PCE, TCE, c-DCE and VC were compiled from the literature. The parameters were used in a casestudy of pump-and-treat remediation of a PCE-contaminated site near Tilburg, The Netherlands. Transport,
non-equilibrium sorption and biodegradation chain processes at the site were simulated using the CHAIN_2D
code without further calibration. The modelled PCE compared reasonably well with observed PCE concentrations in the pumped water. We also performed a scenario analysis by applying several increased reductive
dechlorination rates, re¯ecting dierent degradation conditions (e.g. addition of yeast extract and citrate). The
scenario analysis predicted considerably higher concentrations of the degradation products as a result ofenhanced reductive dechlorination of PCE. The predicted levels of the very toxic compound VC were now anorder of magnitude above the maximum permissible concentration levels. Copyright # 1999 John Wiley &Sons, Ltd.
KEY WORDS

biodegradation; sorption; transport; chlorinated aliphatic hydrocarbons (CAH's); numericalmodelling

INTRODUCTION
Chlorinated aliphatic hydrocarbons (CAH's) are some of the most widespread groundwater contaminants in
the industrialised world (KaÈstner, 1991; Wilson, 1996). Perchloroethylene (PCE) and trichloroethylene(TCE) are used on a large scale as solvents, degreasing agents for metal surfaces and electronics, dry-cleaning
and for a wide range of applications in chemical reactions. Because of spills and accidents, these hydrocarbons have become important contaminants in soil and groundwater. There is a growing concern about
these compounds since they are known to persist in the environment over a long period of time, because of
their xenobiotic origin, and because they are suspected carcinogens, even at low concentrations. Concern
about contamination with PCE or TCE has become more acute, owing to the observation that certain
anaerobic bacteria in groundwater are able to transform these compounds to cis-dichloroethylene (c-DCE)
and vinylchloride (VC). The latter compound is known to be a potent mutagen and carcinogen in humans(Badaracco, 1985). Anaerobic biological transformations of some CAH's can occur within the time frame ofinterest in groundwater (McCarty, 1997). In certain situations, biological degradation of CAH's can be* Correspondence to: J. Schaerlaekens, Laboratory for Soil and Water, Vital De Costerstraat 102, B-3000 Leuven, Belgium. E-mail:jan.schaerlaekens@agr.kuleuven.ac.be
CCC 0885±6087/99/172847±13$1750
Copyright # 1999 John Wiley & Sons, Ltd.

Received 10 September 1998
Revised 16 March 1999
Accepted 8 April 1999

2848

J. SCHAERLAEKENS ET AL.

stimulated by providing optimal redox conditions or by adding nutrients, co-enzymes or electron donors.
This can increase the rate of remediation and decrease the mass of contaminant to be treated (Madsen, 1991;
Fetter, 1992).
Numerical modelling, in addition to laboratory and ®eld-scale experimentation, provides the means toinvestigate the in¯uence of dierent processes on the eectiveness of remediation techniques and the risk ofaccumulation of toxic degradation products. A thorough understanding of the basic processes of transportand fate of CAH's in the environment is necessary in order to reduce the risk to human health, and tosuccessfully formulate and implement remediation techniques.
Most sites contaminated with chlorinated aliphatic hydrocarbons are remediated by pump-and-treat
techniques (EPA, 1997). Numerical models, often embedded in optimization routines, are often used to
predict the eectiveness of and the time needed for such clean-up operations.
In the present study, a numerical code, CHAIN_2D (SÆimuÃnek and van Genuchten, 1994), was used toevaluate the clean-up of a PCE-contaminated shallow aquifer. CHAIN_2D incorporates chemical chain
reactions in classical solute transport equations to make it possible, not only to predict the concentration of
the contaminants themselves, but also the concentration of their reaction products. The decrease in contaminant concentration due to biodegradation becomes especially signi®cant at low pumping rates, andneeds to be incorporated in the model. The objectives of this study were to: (1) compile values for
degradation rate constants, sorption coecients and mass exchange rates from a literature review; (2)compare ®eld observations on PCE degradation under pump-and-treat conditions with model predictions
obtained without extensive model calibration; and (3) numerically illustrate the eect of stimulated
biodegradation on reaction products such as the toxic VC.

THEORY
Reductive dechlorination
Abiotic chemical degradation of PCE occurs very slowly. More important is the microbiologicaltransformation of this chemical. Oxidation of PCE in aerobic environments is unlikely because of the highoxidation grade of the molecule (Vogel et al., 1987). PCE can only be degraded successfully by reductive
processes (Seng, 1991). This reaction occurs preferentially under methanogenic or sulfate-reducing conditions (Wilson, 1996). Such conditions are commonly found in the saturated zone of the soil, or can becreated by adding organic compounds. Microorganisms may be using CAH's in their energy metabolism.
This process involves unique reactions since the microorganisms use CAH's as electron acceptors similar to
the use of oxygen by aerobic organisms. This, in turn, requires suitable electron donors such as sugars or
short aliphatic chains (McCarty, 1997). The anaerobic degradation process also requires, besides a sucientconcentration of electron donors, an optimal redox potential. Other chemical compounds can in¯uence thedegradation of CAH's in dierent ways: they can aect the solubility, and thus the bioavailability, of CAH's;
have a stimulating or toxic in¯uence on microbial growth; serve as a primary substrate for bacteria; in¯uence
redox conditions; or induce the production of enzymes necessary for CAH degradation (Praamstra, 1996).
The most important pathway for anaerobic biodegradation of highly chlorinated CAH's (e.g. PCE, TCE) is
reductive dechlorination (Vogel et al., 1987). The chlorinated ethenes can be reductively dechlorinated in asequential manner, ultimately yielding ethene, which is environmentally acceptable and does not cause direct
health eects (Bosma, 1994; Gosset and Zinder, 1997) (see Figure 1).
In this process, the four chlorine atoms are sequentially replaced by hydrogen atoms in a hydrogenolysis
reaction (Hollinger, 1992). The hydrogen atoms are provided by the substrate that is being oxidized (i.e. theelectron donor). The release of chlorine atoms in water provides energy. A relatively large amount of substrateis necessary for this reaction pathway (Geritse et al., 1995). The rate of degradation of these contaminants
strongly depends on the concentration of the substrate. First-order kinetics is, in most cases, adequate fordescribing reductive dechlorination at relatively low concentrations (Fetter, 1992; Praamstra, 1996).
Copyright # 1999 John Wiley & Sons, Ltd.

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CHLORINATED ALIPHATIC HYDROCARBONS

2849

Figure 1. Reductive dechlorination of chlorinated ethenes under anaerobic conditions

Biostimulation
In some cases, natural degradation processes are sucient to solve the problem of site contamination with
CAH's (Madsen, 1991; Fetter, 1992). However, as the degradation of PCE at contaminated sites often occurs
under relatively unfavourable conditions (i.e. low temperature; low concentration of electron donors,
bacteria and nutrients; low bioavailability of the contaminants), it is often recommended to enhance the rateof biodegradation. This can be achieved in several ways: essential nutrients like phosphate and nitrate can beprovided, co-enzymes can be supplied or electron donors like methanol, methane, glucose, acetate, citrate(e.g. Bagley and Gosset, 1990; Hollinger, 1992; Travis and Rosenberg, 1997) can be added. Buchanan et al.
(1994) used yeast extract (co-enzymes) and citrate (electron donor) to enhance reductive dechlorination of
PCE.
Sorption
Sorption on soil organic matter is an important factor aecting the transport of organic contaminants
during pump-and-treat remediation. Pump-and-treat, however, often creates arti®cially large groundwater¯ow velocities, a situation that may lead to the presence of dierent types of sorption- and diusion-related
non-equilibria (Brusseau and Rao, 1991). Sorption of chlorinated ethenes on soil organic matter is, inessence, a partitioning process between the organic matrix and water. By losing the organic molecule, thewater phase wins free energy (Scow and Johnson, 1997). At high groundwater velocities, desorption from theorganic matrix may become rate limited, leading to kinetic or chemical non-equilibrium (van Genuchten and
Cleary, 1979; Pignatello and Xing, 1996). Since the (de)sorption of contaminants on ( from) soil organic
matter determines their bioavailability (Brusseau, 1995; Scow and Johnson, 1997), their transport characteristics, sorption characteristics and desorption rates are important parameters for solute modelling during
pump-and-treat remediation.
Description of the numerical model
The numerical model CHAIN_2D (SÆimuÃnek and Genuchten, 1994) may be used to numerically simulate
two-dimensional water ¯ow, heat movement and solute transport in saturated and unsaturated soils.
Variably saturated water ¯ow in porous media is based on the Richards equation. Sequential ®rst-order
Copyright # 1999 John Wiley & Sons, Ltd.

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2850

J. SCHAERLAEKENS ET AL.

transformation reactions such as those that occur during sequential dechlorination have been incorporated
in CHAIN_2D. Transport of the dierent components of the degradation chain is described by advection±
dispersion type equations. The equations for the reagent and the reaction products are, respectively:!
@yc1 @rs1@q c@
ij @c10yD1
1ÿ i 1 ÿ m1 m1 yc1
@t
@t
@xj
@xi
@xi
@yck @rsk
@
ij @c
yDk k


@xi
@t
@t
@xj

!
ÿ

@qi ck00
ÿ mk  mk yck  mkÿ1 yckÿ1@xi

2

with k 2 (2, . . . , ns), and where c is the solute concentration in soil±water [M L ÿ3], s the sorbed soluteconcentration [M M ÿ1 soil], r is the bulk density [M L ÿ3], t is time [T], xi are the spatial co-ordinates [L], y is
the water content [L3 L ÿ3], Dij the dispersion coecient tensor [L2 T ÿ1], qi the ith component of the Darcian
velocity [L T ÿ1], m a ®rst-order degradation rate constant [T ÿ1], m0 the ®rst-order rate constant providing
connection between individual species in the chain reaction [T ÿ1], k represents the kth chain number, and ns
is the number of chemical substances in the chain reaction. Although CHAIN_2D includes provisions fortransport in the gas phase assuming linear equilibrium between liquid and gaseous phase, these options were
not considered in the simulations. Degradation assumes that Equations (1) and (2) only occur in the liquid
phase. Scow and Johnson (1997) showed that this hypothesis is, in many circumstances, acceptable.
Degradation is often mimicked by a ®rst-order process:ÿ@ck
 mk ck@t

3

Sorption on the solid phase sk is described in CHAIN_2D by a generalized equation, of the following form:bsk Kd;k ckkb

1  nk c k k

4where Kd,k [L3 M ÿ1] and nk [L M ÿ1] are empirical coecients. The Freundlich, Langmuir and linear
adsorption equations are special cases of Equation (4). In this study we assume nk  0 such that bk  1[Equations (1) and (2) are linear, and Equation (4) reduces to the linearized Freundlich equation]. Chemical
non-equilibrium is described by a two-site approach, to permit consideration of rate-limited desorption. The
two-site sorption concept proposes that sorption sites can be divided into two fractions, sek and snk, following:ens k  sk  sk

5

Sorption on type-1 sites (sek) [M M ÿ1 soil] is assumed to be instantaneous, while sorption on the remaining
type-2 sites (snk) [M M ÿ1 soil] is considered to be time dependent. At equilibrium, the sorbed concentrations
on type-1 (equilibrium) and type-2 (kinetic) sites are, respectively:sk  F k sk

e

6

n

7

sk  1 ÿ Fk skwhere Fk is the fraction of exchange sites in equilibrium with the solution phase [-] and sk is the total amountof sorption sites. Because type-1 sorption sites are always at equilibrium, dierentiation of Equation (6)yields the sorption rate for the type-1 equilibrium sites:@sek@s Fk sk k@t@t
Copyright # 1999 John Wiley & Sons, Ltd.

8

HYDROLOGICAL PROCESSES, VOL. 13, 2847±2859 (1999)

CHLORINATED ALIPHATIC HYDROCARBONS

2851

Sorption on type-2 sites is assumed to be a ®rst-order kinetic process. Dierentiation of Equation (7) gives:"#b
Kd;k ckk
@snk
n
ÿ sk
 ak 1 ÿ Fk 
9b@t1n c kk kwhere ak is the rate constant for adsorption of the kth solute on type-2 sites.
LITERATURE REVIEW ON TRANSPORT AND BIODEGRADATION PARAMETERS
To model transport and degradation of chlorinated ethenes, sorption, mass exchange and degradation
parameters for each chemical in the degradation chain of Figure 1 are required. The parameters characterizing sorption and biodegradation of CAHs were collected from recent literature and are listed inTables I±III. The literature review of parameter values helps to determine a reasonable parameter range.

Table I. Literature values of KOW and estimated values of KOC using the equation log KOC  0.95 log KOW ÿ 0.2(Ball and Roberts, 1991)
Solutelog KOW (-)

KOC (m3/kg)

PCE
PCE
PCE
PCE
PCE
PCE
PCE
PCE
PCE
PCE
PCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
TCE
c-DCE
c-DCE
t-DCE
t-DCE
t-DCE
t-DCE
1,1-DCE
1,1-DCE
1,1-DCE
VC
VC
VC

2.82
2.88
2.60
2.60
2.60
2.53
2.88
2.76
2.61
2.60
2.95
2.50
2.29
2.50
2.42
2.34
2.39
2.20
2.15
2.10
2.15
2.45
2.35
1.48
2.07
2.07
1.88
1.87
1.48
1.48
2.12
2.67
0.6
1.17
0.60

0.303
0.343
0.186
0.186
0.186
0.160
0.343
0.269
0.190
0.186
0.400
0.152
0.138
0.152
0.125
0.132
0.118
0.078
0.071
0.062
0.070
0.137
0.107
0.016
0.059
0.059
0.039
0.038
0.016
0.016
0.065
0.217
0.00230.00820.0023Copyright # 1999 John Wiley & Sons, Ltd.

Reference
Fetter, 1988
EPA, 1979
Ball and Roberts, 1991
Fetter, 1992
Curtis et al., 1986
Banerjee et al., 1980
Schwarzenbach and Westall, 1981
Brusseau et al.,1991
Brusseau et al., 1990
Keizer, 1992
Minister of Supply and Services, 1993aFetter, 1988
EPA, 1979
Banerjee et al., 1980
Schwarzenbach and Westall, 1981
Hutzler et al., 1986
Hutzler et al., 1986
Keizer, 1992
Brusseau, 1992
Brusseau et al., 1991
Brusseau et al., 1991
Brusseau et al., 1991
Minister of Supply and Services, 1993b
EPA, 1979
Rippen, 1984
Rippen, 1984
Fetter, 1988
Brusseau et al., 1991
EPA, 1979
EPA, 1979
Rippen, 1984
Fetter, 1988
EPA, 1979
Rippen, 1984
Schwarzenbach and Westall, 1981
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2852

J. SCHAERLAEKENS ET AL.

Table II. Regression values for the mass transfer rate (log a  D log Kd  E) for a groundwater velocity of 4 cm/h and asoil temperature of 10 8C
Solute
HOC's
TCE
PCE, TCE, t-DCE
PCE, TCE, TCM, DCB, BP,naphthalene, xylene

D

E

r2

Referenceÿ0.668
ÿ0.83
ÿ0.47
ÿ0.70

0.301
0.92
0.56
ÿ1.67

0.95
Ð
0.85
0.80

Brusseau et al., 1989
Brusseau et al., 1991
Brusseau and Rao, 1991
Brusseau et al., 1991

Table III. Published ®rst-order biodegradation rates (m) and half-time values (t1/2) for reductive dechlorination of CAH's
Solutek(d ÿ1)

t1/2 (d) System characteristics

PCE

0.002

347

PCE

0.0075±0.071

PCE
PCE
TCE

0.00790.0021
0.003

9.8
92
88
330
231

TCE

0.001±0.003

TCE

0.008

231
693
87

TCE

0.0062

112

TCE
TCE

0.004
0.0001±0.003

TCE

0.0033

173
231
6930
210

TCE
c-DCE

0.006
0.0014±0.002

c-DCE

0.004

116
347
495
173

t-DCE

0.0017

408

t-DCE

0.007

99

1,1-DCE 0.001±0.0026
1,1-DCE

0.0084267
693
83

1,1-DCE

0.0052133

1,1-DCE

0.006

116

0.0005±0.002

347
1386

VC

Organic carbon
content (%)

Aquifer, sand, anaerobic
conditions
Anaerobic conditions, sulfatereducing conditions
Anaerobic conditions

2.8

Aquifer, sand, anaerobic
conditions
In situ aquifer, anaerobe

2.8

Microcosm, silt/clay/sand,anaerobic conditions
Microcosm loamy sand, sulfate
reducing
In situ, no addition
Microcosm aquifer, anaerobic
conditions
Microcosm aquifer, anaerobic
conditions
Anaerobic conditions
In situ, aquifer, anaerobic
conditions
Microcosm sediment,
anaerobic conditions
In situ, aquifer, anaerobic
conditions
Microcosm sediment,
anaerobic conditions
In situ, aquifer, anaerobic
conditions
Alluvial silt/clay/sand,Anaerobic conditions
Microcosm sediment,
anaerobic conditions
Microcosm sediment,
anaerobic conditions
In situ, aquifer, anaerobic
conditions

Copyright # 1999 John Wiley & Sons, Ltd.

0.25

T
( 8C)

Reference
Roberts et al., 1982

35
35

Bagley and Gosset,
1990
Bouwer et al., 1981
Praamstra, 1996
Roberts et al., 1982

10

Wilson et al., 1994

0.34

17

Wilson et al., 1996

2

25

Barrio-Lage et al., 1987

1

Poulsen et al., 1996
Martin andImbrigiotta, 1994
Kleopfer et al., 1985
35
10

Bouwer et al., 1981
Wilson et al., 1994

25

Barrio-Lage et al., 1986
Wilson et al., 1994

Wilson et al., 1994
Wilson, 1996
Barrio-Lage et al., 1986
Barrio-Lage et al., 1986
Wilson et al., 1994

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CHLORINATED ALIPHATIC HYDROCARBONS

2853

Soil water partitioning coecients
Organic chemicals generally are subject to partitioning between the liquid phase and the solid phase(mostly on soil organic matter). Partitioning between the liquid and solid phases in soils with organic carbon
content 40.1% can be assessed from the organic carbon±water partitioning coecient, KOC [L3 M ÿ1](Harmon and Roberts, 1994):
Kd  KOC *fOC

10

where Kd [L3 M ÿ1] is the soil±water partitioning coecient, and fOC [-] is the fraction of organic carbon of
the soil. The value of KOC , in turn, can be assessed from KOW [-], the octanol±water partitioning coecient.
The value of KOW re¯ects the hydrophobic characteristics of the molecule and can be assessed either from thechemical structure or can be measured. An empirical correlation between KOC and KOW is (Scow andJohnson, 1997):
log KOC  A log KOW  B

11

where A and B are regression coecients. Ball and Roberts (1991) and Schwarzenbach and Westall (1981)
give a review of regression values for A and B. The value of A was mostly in a range of 0.9±1.0, whereas
B ranged between ÿ0.3 and 0.4. Previously published KOC values are listed in Table I.
Fraction of type 1 sites
The fraction of type-1 sites (equilibrium) was similarly estimated by means of a regression equation.
Brusseau et al. (1991) determined the regression equation for CAH's as:
Fk  0102 log KOW  C

12with r2  0.80. The value of Fk was found to change slightly with groundwater velocity (Brusseau et al.,1991); for chlorinated ethenes the value of F varied between 0.4 and 0.6 (Brusseau et al., 1991; Brusseau,
1992). We found that CHAIN_2D predictions were not very sensitive to this parameter for given adsorption
rates (Schaerlaekens, 1997).
First-order mass transfer rate
Several empirical relationships exist for the correlation between the soil±water partitioning coecient andthe mass transfer coecients for given values of the pore±water velocity (Brusseau and Rao, 1991; Brusseau
et al., 1991; Brusseau, 1992):
log a  D log Kd  E

13

where D and E are regression coecients. A summary of regression values from the literature is given in
Table III for a ¯ow velocity of 4 cm h ÿ1 and a soil temperature of 10 8C.
Biodegradation rate
A wide range of values was found for biodegradation rates of CAH's. The values are very site speci®c anddier as a function of oxidation potential, nutrients present, concentration of contaminants, and presence of
microrganisms, among other factors. Whereas dierent degradation rate equations exist, we assumed that®rst-order degradation is a reasonable assumption for biological degradation involving electron acceptors at
low concentrations (Praamstra, 1996). Very few ®eld data were available for biodegradation rates, and we
obtained most data from literature microcosm experiments. If no values were given, the rate was calculated
from the published half-life times. A review for parameters obtained in this manner is given in Table III.
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J. SCHAERLAEKENS ET AL.

Table IV. Measured initial soil±water concentrations as
a function of radial distance to extraction wells (m)cPCE (mg/l)

4.3
8.6
17.2
23.5
30.1
33.0
38.1
44.0
53.9

26.00
16.25
9.75
6.50
5.20
1.95
0.65
0.19
Ð

CASE STUDY
Field site
The CHAIN_2D code was applied to a heavily polluted site near the town of Tilburg, The Netherlands.
The site contained many chlorinated hydrocarbons, of which 65% consisted of PCE and 30% of
1,1,1-tricloroethane (TCA). Degradation products of these contaminants were below the detection limits.
The site was previously characterized and described by Praamstra (1996) and served the purpose ofevaluating CHAIN_2D calculations. The 13-m thick top layer of the aquifer consisted of a sandy,
Pleistocene material, with 0.44% organic matter and several loam and peat lenses. The groundwater levelwas 1 m below the surface. The regional ¯ow was negligible. CAH contamination was found throughout thewhole saturated zone of the aquifer and reached the extremely low permeability bottom clay layer. A threedimensional characterization of the contaminant plume was based on Praamstra (1996). PCE concentrationsof the contaminants were measured as a function of the radial distance to the extraction well. Results areshown in Table IV. Decay products of PCE, that were probably present at low concentrations owing todegradation were not detected at the site.
A single-well pump-and-treat scheme was installed over the full depth of the saturated zone and pumping
was started on 14 July 1994, with an average extraction rate of 144 m3 d ÿ1. The pump was disconnected
twice, from 12±18 December 1994 and from 13±28 March 1995. Chemical analyses of PCE concentrations in
the pumped water were carried out monthly.
MODEL IMPLEMENTATION
A 110-m diameter, circular computational domain, with the pumping well at its centre, was generated using
the unstructured mesh generator provided in HYDRUS_2D (SÆimuÃnek et al., 1996). At each node a set of¯ow and transport parameters, based on the properties of the present soil type and the measured initial PCE
concentrations, were assigned. No decay products were detected at the site. Modelling the site's history inorder to estimate these initial concentrations would require an extended knowledge of the history of thecontamination and of the (bio)chemical properties of the soil, which were not available. Site-speci®c values
for chemical transport parameters, based on laboratory measurements, were not available. Therefore, allparameters were estimated from the values and relationships discussed in the literature review. Table V lists
chemical transport and degradation parameters for PCE and its daughter products as used for thesimulation.
The soil±water partitioning coecients were calculated from the organic carbon±water partitioningcoecients by applying Equation (10) using the site-speci®c fraction of soil organic matter, withfOC  0.004. The latter was calculated from the octanol±water partitioning coecients of the dierent
chemicals. The degradation rates for PCE and TCE were taken from Roberts et al. (1982) whereas those for
Copyright # 1999 John Wiley & Sons, Ltd.

HYDROLOGICAL PROCESSES, VOL. 13, 2847±2859 (1999)

2855

CHLORINATED ALIPHATIC HYDROCARBONS

Table V. Chemical transport and biodegradation input parameters for the CHAIN_2D model. Ks denotes the saturated
conductivity, r the bulk density, Fk the fraction of type 1 sites, Kd the partition coecient soil±water, m the ®rst-orderdegradation rate and a the mass transfer rate
Parameter

Soil

Ks (m/s)r (kg/m3)
Fk (ÿ)Kd (m3/kg)
a (d ÿ1)m (d ÿ1) without biostimulationm (s ÿ1) in case of biostimulation

8.10  10
1550
0.45

PCE

TCE

c-DCE

VC

784
0.002
0.005
3.82  10 ÿ7

277
0.003
0.005
2.43  10 ÿ71530.00140.005
2.20  10 ÿ7

10.6
0.00050.005
5.78  10 ÿ8ÿ5c-DCE and VC were taken from Wilson et al. (1994). The ®rst-order rate constants for desorption fromtype-2 sorption sites, presumably mainly organic matter (Pignatello and Xing, 1996), were obtained from
Brusseau (1992). The fraction of type-1 sites was taken from Praamstra (1996).
MODELLING RESULTS
Figure 2 shows numerical calculations of the concentration of PCE and its degradation products in theextraction water, from the start of the pumping (t  0) until 1000 days later. Measured PCE concentrations
are also indicated. The measured reaction products never reached concentrations above the detection limit of
1 mg l ÿ1. Concentrations of simulated decay products were never higher than this detection limit. Since noFigure 2. Simulated concentrations of PCE and its reaction products in pumping water
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2856

J. SCHAERLAEKENS ET AL.

decay products were detected at the start of the pump-and-treat remediation, a zero concentration initialcondition was applied for all decay products, although it is reasonable to suppose non-zero concentrations ofTCE, c-DCE and VC owing to degradation. Lack of information on the history of the contamination(including details on the source term) did not allow them to be estimated by numerical simulation. Theselimitations to the numerical study, however, do not aect the predicted PCE concentrations.
In spite of the fact that all physicochemical parameters were estimated from the literature, the numericalsimulation follows the measured values quite well, except for the last measured value of PCE, which was
almost three times lower than the calculated value. The somewhat lower measured PCE values, certainly of
the last data point, could possibly be the result of volatilization of PCE from the sampled well water, since
PCE is known to have a relatively high Henry's constant. The remediation was terminated after about
250 days because remediation objectives were judged to be met (Praamstra, 1996). The calculated concentrations of the reaction products in soil and groundwater were very low and of no harm to public health. TCE
and DCE have almost the same soil±water partitioning coecient as PCE, and are extracted at relatively thesame rate as PCE. VC on the other hand has a substantially lower partitioning coecient and is removed
much faster. Since no initial concentrations were known for these degradation substances, it is dicult tomake a good estimate of their concentrations in the initial phase of the pumping.
For the biostimulation scenario, the degradation rates of Buchanan et al. (1994) were used (Table V).
Their data were based on in situ tests in which yeast extract and citrate were added to an aquifer in order tostimulate growth of methanotrophic bacteria, and thus enhance PCE degradation. The degradation rates forthe dierent chlorinated ethenes in the case of biostimulation were considerably higher than without
biostimulation. Figure 3 shows the results of the scenario analysis with the higher degradation rates. The

Figure 3. Simulated concentrations of PCE and its reaction products in a biostimulation scenario
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CHLORINATED ALIPHATIC HYDROCARBONS

2857

total concentration of chlorinated ethenes in the extraction water declined much faster than for the scenario
without stimulation.
However, the reaction products now reached higher concentrations than in the standard scenario, andexceeded the groundwater standards. Since the concentrations obtained for these chemicals are higher thanthe detection limit, the zero concentration initial condition has no major in¯uence on the conclusions for this
scenario; only the absolute values could have been higher. Since VC is much more hazardous than PCE, thepumping had to continue until the concentration of the reaction products reached acceptable values. The
total remediation time, however, decreased, compared with the standard scenario.
CONCLUSIONS
Pump-and-treat remediation technologies have often been applied without considering a detailed design,
based on numerical modelling, in which the physical, biological and chemical processes are represented as
best as possible. Lack of detailed process models and parameters certainly contributed to this practice. In
this paper a potentially useful model for the optimal design of pump-and-treat remediation is presented andapplied to a PCE-contaminated site in the Netherlands. The most important parameters controlling sorption
and rate-limited mass transfer between the liquid and organic matter, and the degradation rate for the chain
reaction PCE ! TCE ! c-DCE ! VC were obtained from a compilation of literature values. Given the
scarcity of site-speci®c model parameters, simulations with CHAIN_2D of PCE removal from the aquifer
were in relatively good agreement with PCE measurements. Observations and simulations showed that, inthe absence of biostimulation, the concentration of PCE degradation products remained below the detection
limit. Because the soil±water partitioning coecients for TCE and c-DCE were of the same order as that for
PCE, removal of TCE and c-DCE occurred at relatively the same rate as PCE removal. The toxic component, VC, had a much lower soil±water partitioning coecient compared with that of PCE, and its
removal occurred at a much faster rate (assuming the same mass transfer parameter).
When the biodegradation rates of in situ biostimulation conditions were simulated, PCE degradation
products reached their maximum permissible level much faster than without biodegradation enhancement.
The concentrations of the degradation products TCE, c-DCE and VC were now close to, respectively one,
two and three orders of magnitude higher compared with the concentrations obtained in the absence ofbiostimulation. VC temporarily reached critical concentrations above the maximum permissible contaminant level. In the case of biostimulation it is important to closely monitor any toxic degradation products, inaddition to the main contaminant; unforeseen high levels of toxic degradation products may otherwise go
undetected.
LIST OF SYMBOLS
csolute concentration in soil water [M L ÿ3]

Dij

dispersion tensor [L2 T ÿ1]

fOC

fraction of organic carbon [-]Fkfraction of type 1 sites [-]ksubscript of the nth step in the chain reaction

KOC organic carbon±water partitioning coecient [L3 M ÿ1]
KOW octanol±water partitioning coecient [-]Kd

soil±water partitioning constant [L3 M ÿ1]

Ks

saturated conductivity [L T ÿ1]ns

number of compounds

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HYDROLOGICAL PROCESSES, VOL. 13, 2847±2859 (1999)

2858

J. SCHAERLAEKENS ET AL.

q

Darcy velocity [L T ÿ1]ssolute concentration sorbed on to the soils [M M ÿ1 soil]sesolute concentration sorbed on type 1 sites [M M ÿ1 soil]

snsolute concentration sorbed on type 2 sites [M M ÿ1 soil]stotal solute concentration sorbed on the soil [M M ÿ1 soil]ttime [T]x

distance [L]a®rst-order mass transfer coecient (T ÿ1)

bk

Freundlich experiment (-)

m®rst-order degradation rate (T ÿ1)

m0®rst-order production rate (T ÿ1)rbulk density (M L ÿ3)yporosity or water constant at saturation (L3/L3)

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