2009; Foster, 2001). Today, aquifers are increasingly being
used for aquifer storage and recovery (Dillon et al., 2006; Pyne,
2005), facilitated storm water infiltration (Ferguson, 1990),
storage (Steeneveldt et al., 2006), and brine disposal
(Tsang et al., 2008) and are a target for remediation of
contamination (European Environment Agency, 2007). During
the past decade, a rising demand for sustainable energy
sources has led to intensified use of seasonal aquifer thermal
energy storage (SATES), a cost-effective energy technology in
support of ambitions for CO
-emission reductions. This
technology provides seasonal heating and cooling of buildings
by means of the alternating injection and recovery of heated
(injection during summer, recovery during winter) and cooled
(injection during winter, recovery during summer) ground-
water via wells in aquifers (Carotenuto et al., 1991; Edworthy
and Puri, 1986; Kim et al., 2010; Molz et al., 1978, 1979, 1981;
Tsang and Hopkins, 1982). Most of those SATES systems
are operated with only limited temperature differences
(ΔTb15 °C) between warm (b20 °C) and cold wells (~5 °C)
in shallow aquifers with an ambient groundwater temperature
of ~11 °C. An increasing number of SATES systems is reported
in European countries and elsewhere (Gao et al., 2009; Sanner
et al., 2003). In the Netherlands, for example, it is assumed that
SATES will be the largest form of groundwater usage by 2020,
pumping 1225 to 6300 million m
per year, thereby probably
exceeding the total drinking-, industrial-, and agricultural
groundwater extraction of 1500 million m
per year (Bonte et
Depending on their location, spatial density, and capacity,
the associated rise in groundwater use by these SATES
systems may cause interference between SATES systems
and groundwater extractions (Bonte et al., 2011a,b). SATES is
also likely to affect soil and groundwater contamination,
since the largest demands for heat and cold exist in
urbanized and industrial areas. The depth at which SATES
systems are typically operated (>10 to b250 m below the
water table) coincides with the depth at which groundwater
contaminant plumes of chlorinated hydrocarbons (CHCs)
are generally present, particularly tetrachloroethene (PCE)
and trichloroethene (TCE) and their daughter products
dichloroethene (DCE) and vinyl chloride (VC). Such CHC
plumes develop through the long-term dissolution of dense
non-aqueous phase liquids (DNAPLs) often used in dry
cleaning and metal degreasing processes and are one of the
most prevalent organic contaminations in urban groundwa-
ter (e.g., Bradley, 2000; Parker et al., 2003; Smidt and de Vos,
2004; Wiedemeier et al., 1999).
The presence of CHC plumes currently hampers the
development of SATES in urbanized areas in the Netherlands
(Slenders et al., 2010; Taskforce ATES, 2009). But besides
potential negative effects associated with pumping of contam-
inated groundwater, several potentially positive effects of
SATES systems on the transport and degradation of contam-
ination plumes have been suggested. These presume the
anticipated stimulation of contaminant degradation due to
the effect of periodically elevated temperatures and enhanced
mixing of microorganisms, nutrients, oxidants, reductants, and
contaminants within SATES systems (Slenders et al., 2010;
Verburg et al., 2010). Since temperature differences in current
SATES systems are relatively limited and pumped volumes are
large, fundamental processes like spreading of (diluted)
contaminants and enhanced DNAPL dissolution increasing
dissolved contaminant mass need special consideration.
However, they have so far not been evaluated in the
assessment of the impacts of SATES on CHC contaminated
aquifers and focus has been mainly on modeling of temper-
ature and carbonate precipitation in relation to well clogging
(e.g., Brons et al., 1991; Molz et al., 1983). This knowledge gap
in water quality development has been a severe constraint in
SATES application in many urban aquifers.
Here, we present a first modeling approach to study the
effects of alternating transient pumping by low-temperature
SATES systems on CHC contaminated aquifers. The aims of
this study are to:
▪assess the extent to which various factors (hydrogeology,
contaminant degradation kinetics and DNAPL dissolution)
can affect CHC plumes when SATES is applied;
▪evaluate the conditions under which SATES systems are
likely to have either positive or negative effects on CHC
plumes in the presence or absence of DNAPLs within the
zone of influence of SATES wells.
Temperature dependencies could be ignored due to the
limited temperature differences of current thermally bal-
anced SATES systems and were therefore not incorporated in
the model. This study underlines potential benefits and
identifies the risks of SATES application in CHC contaminated
aquifers, prior to the approval of field implementation. An
existing SATES system with therefore a hypothetical PCE
aquifer contamination was modeled with a customized
version of PHT3D.
2. Material and methods
2.1. Model components
In this study, the reactive multi-component transport
model for saturated porous media PHT3D Version 2
(Prommer and Post, 2010; Prommer et al., 2003) was used
for the simulation of groundwater flow, solute transport, and
degradation of contaminants (Fig. 1). This coupled model of
MODFLOW/MT3DMS (Harbough et al., 2000; Zheng and
Wang, 1999) and PHREEQC-2 (Parkhurst and Appelo, 1999)
enables the simulation of advection, dispersion, diffusion,
and almost any desired chemical reaction. In this study, the
processes of DNAPL dissolution and sequential degradation of
dissolved CHCs were included in a three-dimensional flow
field. The MT3DMS-based version of PHT3D does not allow
simulating the effects of density and viscosity changes due to
the effect of temperature differences in the SATES system on
groundwater flow. Therefore, a version based on SEAWAT
version 4 (Langevin et al., 2008) was used for this purpose,
which was coupled with PHT3D Version 2 in the same way as
was done previously by others (Mao et al., 2006; Post and
Prommer, 2007; Robinson et al., 2009). Temperature effects
were simulated initially, but did not affect the outcomes of
the initial stage of modeling. Therefore, temperature differ-
ences were not subsequently considered.
In order to realistically simulate the distribution of
injected and recovered water over several screened intervals
of the wells, as well as the mixing of water qualities in the
wells during recovery from several model layers by one well,
2K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
the multi-node-well (MNW) Version 1 package (Halford et
al., 2002) was used. A modification was made to the code in
order to enable recirculation of the recovered mixed water in
the SATES system to the injection well. This modification
enabled recirculation for the MNW in the same way as for the
existing recirculation for a single-node well (Zheng, 2006).
2.2. Reaction network
2.2.1. Degradation of CHCs
In this study the full degradation sequence from PCE to
ethene as described by Wiedemeier et al. (1999) was
modeled. The kinetics of this process have been described
by various rate models, mainly zero-order, first-order with
respect to contaminant concentration, and Michaelis–
Menten type (Bekins et al., 1998; Suarez and Rifai, 1999).
Depending on which factor limits the overall degradation
rate, the applicable type of kinetic rate expression for the
degradation of CHCs varies, thereby affecting overall degra-
dation within SATES systems. In our study we applied
zero-order degradation kinetics for the cases where degra-
dation rates were limited by the slow release of electron
donor from the sediment (e.g., McCarty, 1997; Pavlostathis
and Zhuang, 1993), i.e. by hydrolysis of sedimentary organic
matter (SOM: Fig. 2), typical for most natural aquifer
environments (Hartog et al., 2005; Jakobsen and Postma,
1994; van Helvoort et al., 2007). These simulations will be
referred to as “sediment-limited”(SL) model runs. In the
SL-scenario it is assumed that organic matter within the
model domain degrades at a constant rate in space and time,
independent of CHC concentrations. In contrast, for condi-
tions with sufficient available reactive dissolved organic
matter, degradation kinetics are described as first-order
with respect to contaminant concentrations (Suarez and
Rifai, 1999). These simulations will be referred to as
“contaminant-limited”(CL) model runs (Fig. 2).
2.2.2. Dissolution from DNAPL and adsorption
For the dissolution of PCE from the DNAPL in the aquifer,
the following rate-limited expression was used to describe
Fig. 1. Modeling scheme illustrating input data and model codes used to create simulation output. Mixed concentrations during recovery were calculated
(‘recovered mean concentration’) and provided the input for the injection well.
Fig. 2. Controlling factors assumed for CHC degradation in this study. SOM is the sedimentary organic matter, C
is the concentration of dissolved organic
matter (M L
), t is time (T), λ
is the degradation constant (M L
is the concentration of chlorinated hydrocarbon species considered (M L
is the degradation rate constant for first-order degradation (T
3K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
the mass transfer rate between the immobile DNAPL phase
and the dissolved phase as (e.g., Hayden et al., 1992; Miller et
al., 1990; Prommer et al., 1999):
is the mass of DNAPL PCE per unit pore volume
is a lumped rate-transfer coefficient (T
is the solubility of PCE (M L
) and C
is the concentration of
PCE in the groundwater (M L
). To enhance dissolution and
thereby emphasize the effect of the presence of the DNAPL, no
permeability decrease by the presence of the DNAPL was
modeled, while a relatively high rate-transfer coefficient
=1) was chosen, which is in line with Park and Parker
(2005), resulting in equilibrium dissolution at solubility
concentrations. Adsorption of dissolved CHC species was
modeled using a linear sorption isotherm.
2.3. Case study
As the application of SATES in contaminated aquifers is
restricted to date, no field data exist that can be used for the
purpose of this study. The well-monitored existing SATES
system ‘Uithof’in an uncontaminated aquifer in the city of
Utrecht (The Netherlands) was therefore used as the basis for
the simulations (Figs. 3 and 4). The local hydrogeology and
one pair of wells of the SATES system were represented in the
model. A fictitious contaminant source was added in the
model by placing a DNAPL at the base of the aquifer, 10 m
from one of the SATES wells (Fig. 4,Section 2.5).
2.3.1. Hydrogeological setting
The hydrogeological schematization of the field site
was derived from the regional geological model REGISII.1
(TNO-NITG, 2009). At the ‘Uithof’location (Figs. 3 and 4),
the subsurface consists of unconsolidated deltaic and fluvial
sediments forming alternating sand and/or gravel layers
(aquifers) and clay layers (aquitards). The SATES system is
injecting cold water in the western well (well “C”in Fig. 4) and
warm water in the eastern well (well “W”). This ~47 m thick
aquifer with a transmissivity of 1320 m
/d is covered by a
2.8 m thick Holocene clay layer (K
=0.01 m/d) and underlain
by a 6 m thick aquitard (K
=0.005 m/d). In this aquifer, the
regional groundwater flow is determined by topographically
Fig. 3. Location and hydrogeological setting of the Uithof case study area. I —Holocene clay cover; II —aquifer 1; III —aquitard 1; IV —aquifer 2; V —local clay
layer; VI —aquitard 2; VII —ice pushed ridge; VIII —SATES system ‘Uithof’. Local hydrogeology based on the REGIS II.1 hydrogeological model (TNO-NITG, 2009).
Well separation not to scale.
4K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
elevated ice pushed ridges in the northeast of the study area
(Fig. 3), resulting in a southwestern flow of ~7 m/year at the
SATES location. This estimate is based on a measured hydraulic
gradient of 0.00024 m/m (Faneca-Sànchez et al., 2010), a
porosity of 0.35, and a hydraulic conductivity of 28 m/d.
2.3.2. SATES conﬁguration
The wells of the SATES system are screened across two
stratigraphic sandy sections of high permeability within
aquifer 1 which are separated by a sandy clay layer (Fig. 4).
The weekly injected volumes in these layers were registered
in 2009 (Fig. 5), of which the last 4 weeks were prefixed to
the front of the data set in order to start with a full cold cycle.
A total of 147,951 m
of groundwater was pumped from the
warm well to the cold well during winter, while 132,232 m
was pumped from the cold to the warm well during summer.
Due to the climate control requirements of the building, heat
demand in winter was of shorter duration but more intense
than the cold demand in summer, resulting in this volumetric
2.4. Model set-up
2.4.1. Grid design
The local hydrogeology and SATES configuration in Fig. 4
were incorporated into the PHT3D model. For this model a fine
vertical discretization (∆z= 0.3 m) was used where the DNAPL
was present (at the base of aquifer 1) to allow for steep CHC
concentration gradients in this zone. Since larger gradients of
the CHC concentrations and the temperature were also
expected near both SATES wells, refinement was also made in
the grid near the wells (∆x=3 m, ∆y=0.5 m), whereas the
cells away from the DNAPL and the SATES wells were enlarged
stepwise to a maximum horizontal size of 120 m× 80 m.
The total model extent was 1760 m (west–east)×1323 m
(north–south), and the minimum distance between the wells
and the nearest edge of the model was 660 m, which was
sufficient to prevent boundary effects from affecting the
simulated flow regime and to prevent any contaminant mass
from leaving the system.
2.4.2. Groundwater ﬂow and transport parameters
The values of Table 1 were used for the modeling of
groundwater flow and solute transport. The third-order
total-variation-diminishing (TVD) scheme (Leonard, 1988)
was used since it is mass-conservative and provided the most
stable breakthrough curves at the wells during test runs. The
distribution coefficients for the linear isotherm sorption were
based on the K
(Karickhoff, 1981) for each species and an
organic carbon fraction (f
) of 0.005. This resulted in
retardation factors of 7.2, 2.5, 2.2 and 1.04 in aquifer 1 for
PCE, TCE, DCE and VC, respectively. The solubility of PCE was
set to 1.3 mmol/L (215.6 mg L
Regional flow at the location of the SATES system was
simulated by specifying interpolated regional hydraulic heads
(Faneca-Sànchez et al., 2010) as constant head boundaries at
the edges of the model. These interpolated heads were also
used for the initial hydraulic heads. Preliminary simulations
Fig. 4. Local hydrogeology and SATES configuration. In winter (situation in this figure), cold water is being injected at well ‘C’and warm water is recovered at well
‘W’. In summer, the pumping directions are reversed.
5K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
showed that groundwater flow was not significantly
influenced in the temperature range of the SATES system
simulated (8 °C in the cold bubble to 16 °C in the warm
bubble), and therefore the density and viscosity corrections
were not performed in the final simulations in order to
constrain model runtimes. This approach is supported by Ma
and Zheng (2010), who demonstrated that temperature
effects on fluid flow parameters may be neglected when the
maximum temperature difference is within 15 °C.
A total of 988 stress periods was used, each lasting 7 or
8 d (51 and 1 stress period(s) per year, respectively),
resulting in a total period of 19 years. Within each stress
period, the reaction time step (i.e., a call to PHREEQC using
the concentrations in each cell after transport) was 1 d. The
daily pumped volumes varied per stress period, based on the
measured pumping volumes (Fig. 5).
2.4.3. Degradation rate constants
For this study it was assumed that the CHC degradation
rate is either (i) zero-order sediment-limited (SL) or
(ii) first-order with respect to contaminant concentrations
(contaminant limited: CL). The selected constants for
first-order degradation scenarios (Table 2) are consistent
with those of Suarez and Rifai (1999) and Van Breukelen et
al. (2005). Zero-order degradation rates were less available
but were scaled to the first-order degradation rates. Based on
the degradation constants used, first-order PCE degradation
rates equal zero-order degradation rates at a PCE concentra-
tion of 0.37 mg L
, above which first-order degradation
rates are higher. Reaction rates are known to increase
exponentially with temperature (Arrhenius, 1889), and this
has also been shown for microbial dechlorination rates (e.g.,
Friis et al., 2007). However, for a thermally balanced SATES
system, the temperature increase for groundwater in the
warm bubble is balanced by the temperature decrease in the
cold bubble. The expected rate increase in the warm bubble is
therefore counteracted by a rate decrease in the cold bubble.
Although there will be a net increase of the overall rate due to
the exponential dependence of rates on temperature, this
effect is very small (b1%) for the temperature differences
(ΔTb15 °C) between warm and cold bubbles under which
SATES systems are generally operated, as calculated using the
Arrhenius equation (Hartog, 2011). For this reason, the
temperature dependency of degradation rate constants was
neglected in the model.
2.5. Contamination scenarios
A modeled dissolved PCE plume, which was generated
during 40 years of DNAPL dissolution without further PCE
degradation, was used as the initial contaminant condition for
the four SATES scenarios (Table 3). DNAPL is absent in
Fig. 5. Daily mean air temperature at the weather station De Bilt (~2 km
from the ‘Uithof’SATES system) of the Royal Netherlands Meteorological
Institute (KNMI, 2011) and corresponding injection rates. Week 1 represents
the first week of December, which marked the start of the cold period when
heating was required and cold water was injected.
Set of parameters used in model.
Flow and physical transport parameters Value Unit
Model dimensions l
1760× 1323× 135 m
Discretisation ∆x3–120 m
Discretisation ∆y0.5–80 m
Discretisation ∆z0.3–79 m
Horizontal hydraulic conductivity K
0.1–35 m d
Vertical hydraulic conductivity K
0.01–17.5 m d
Effective porosity η0.35
Specific yield 0.25
Specific storage 0.0001 m
Longitudinal dispersivity α
Horizontal transversal dispersivity α
Vertical transversal dispersivity α
Molecular diffusion coefficient D
Distribution coefficient K′
1.26/ 0.32/ 0.24/ 0.01
Bulk density ρ
1430/ 1573/ 1716 kg m
PCE: perchloroethene, TCE: trichloroethene, DCE: dichloroethene, VC: vinyl
Degradation rate constants.
Species Sediment-limited (SL)
PCE 1.86 0.5
TCE 1.74 0.5
DCE 1.86 0.25
Vinyl chloride 0.11 0.125
6K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
scenarios 1 and 2. In scenarios 3 and 4, a fictitious (PCE)
DNAPL initial source of 630 L (3 ×2× 0.3 m; 1022 kg) was
positioned at the base of aquifer 1 (Fig. 4). Each of these four
model scenarios considered different combinations of specific
factors: the SATES system, occurrence of CHC degradation, and
presence of DNAPL (Table 3). All scenarios were performed for
SL- as well as CL-limited degradation conditions.
2.6. Model output
The development of the PCE plume volume and total mass
(sum of the dissolved and adsorbed mass) in the aquifer and
remaining DNAPL mass were monitored for each scenario
based on the mass budgets reported by PHT3D. In addition,
the total volume of groundwater exceeding Dutch national
groundwater quality criteria was calculated based on the
modeled concentrations, porosity, and the volume of each
grid cell. The desired target contaminant level (TCL) for PCE is
0.01 μg/L and the maximum contaminant level (MCL, above
which remediation may be required) is 40 μg/L (VROM,
The results are described for the scenarios presented in
Table 3. All scenarios started with a PCE plume developed from
the DNAPL near well W with a total plume mass of 29 kg. In
this initial PCE plume, 13.0 Mm
of groundwater exceeded the
national target contaminant level (TCL; 0.01 μg/L PCE), of
which 2.4 Mm
also exceeded the maximum contaminant
level (MCL; 40 μg/L PCE).
3.1. Scenario 1: reference scenario: plume with degradation
For the sediment-limited degradation (SL, zero-order)
scenario, the combined effect of degradation and dispersion
led to a decrease of more than 0.6 Mm
of the plume volume
in which PCE concentrations exceeded the TCL after 19 years.
In this plume, slightly less than 0.6 Mm
also exceeded the
MCL (Fig. 6A). Degradation caused the total PCE plume mass
to decrease by 32.5% (Fig. 7A). These results indicate that
without the SATES system the contaminant plume would
remain present in the aquifer for over several decades.
For the contaminant-limited degradation scenario (CL, first-
order), the decrease of PCE plume mass was 94.6% (Fig. 7B).
Despite this larger overall degradation than in the SL case, the
groundwater volumes exceeding the TCL and MCL for PCE were
respectively larger and comparable to SL degradation (Fig. 6B).
Despite the significant plume mass reduction, the decreasing
degradation rates with decreasing concentrations for this case
mean that several decades are still required to meet the
regulatory concentration limits.
Scenarios analyzed in the case study.
Scenario Characteristics Degradation SATES DNAPL
S-1 Degradation present, no SATES
system and no DNAPL
S-2 Addition of SATES system in
absence of DNAPL
X X –
S-3 SATES and DNAPL present, no
S-4 DNAPL and SATES, present,
X X X
Fig. 6. Total PCE plume volume (top values indicate C
>TCL) for each scenario in case of A) sediment-limited degradation and B) contaminant-limited
degradation. Bars give the volumes only exceeding the TCL (0.01 bC
b40 μg/L) and volumes exceeding the MCL (C
>40 μg/L) after 19 years.
7K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
3.2. Scenario 2: plume capture within the SATES system with
In the case of SL degradation, the introduction of the SATES
system resulted in a rapid decrease of the PCE mass in the
plume. A nearly complete reduction of PCE plume mass of
99.1% was achieved after 19 years, reflecting the effect of
increased degradation due to spreading of the contaminants
across a largersediment volume. A concomitant decrease in the
volume of groundwater exceeding the regulatory limits for PCE
was observed: only 0.1 Mm
of groundwater still exceeded the
MCL, and an even smaller volume exceeded only the TCL.
Daughter products were largely absent after 19 years in this
scenario. VC for instance, reached its peak in plume mass (3 g)
at t=78 d, which decreased to a total mass of 9 mg (Fig. 8A)
and a maximum concentration of 0.07 μg/L, after 19 years.
Contrary to the case with SL degradation, the PCE plume
mass in the case of CL degradation was equal to the situation
without the SATES system (Fig. 7B, scenarios 1 and 2). This
was because the first-order rate expression with respect to
contaminant concentration did not result in an increased PCE
mass removal by spreading over a larger aquifer volume. The
increased contaminated volume was accompanied by de-
creased degradation rates due to the proportional dilution of
concentrations. Moreover, with the first-order degradation
kinetics for the CL case, concentrations in the center of the
plume were lowered more rapidly due to a high degradation
rate, whereas in the reinjected (and therefore diluted) SATES
water and in groundwater at the plume fringe low degradation
rates led tolonger persistence of concentrations above the TCL.
The latter is shown by the PCE plume volume exceeding TCL
especially in scenario 2 (Fig. 6B), which was more than 20,000
Fig. 7. Plume mass development in the PCE plume per scenario during A) sediment-limited (SL) degradation and B) contaminant-limited (CL) degradation.
Seasonal changes in plume mass are caused by SATES pumping scheme, further explained in Section 3.3. The vertical arrows indicate the direction in which the
curves for scenario 4 would shift if the rate constant would be increased or decreased.
Fig. 8. Plume mass development in the VC plume per scenario during A) sediment-limited (SL) degradation and B) contaminant-limited (CL) degradation.
8K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
times larger in this case, compared with the SL degradation
case. VC reached its maximum plume mass of 390 g at the end
of the simulation (Fig. 8B), which is a significantly larger mass
compared to the SL case.
3.3. Scenario 3: presence of DNAPL in absence of degradation
within the SATES system
In this scenario, an increase in PCE plume mass was
caused by enhanced dissolution of the DNAPL. This is due to
the dilution of PCE by mixing in the well during extraction
phases, which allows additional DNAPL to dissolve when the
water with diluted PCE concentrations passes the DNAPL
during the subsequent injection phases. A yearly fluctuation,
caused by the alternating recovery and injection and varying
pumping rates, is superimposed on a nearly linear increase in
PCE plume mass (Fig. 7A,B; scenario 3). This linear increase
indicates that the dilution that occurred by dispersion and
particularly by mixing in the abstraction well was sufficient
to maintain large concentration differences (C
the DNAPL, which was confirmed by concentration observa-
tions in the model. After 19 years, 560 kg of PCE was found in
the groundwater, present in a groundwater volume of
exceeding the TCL, and a volume of 254.5 Mm
even exceeding the MCL (Fig. 9). This indicates a large
increase in contaminated aquifer volume, compared to the
same scenario without SATES introduced, in which 20.2 Mm
and 4.8 Mm
exceeded the TCL and MCL for PCE, respectively.
On the other hand, only 48% of the original DNAPL mass
remained, which means that almost 532 kg of PCE was
dissolved, indicating a mean mass discharge from the DNAPL
of 76.6 g/d. Without SATES, an additional PCE mass of only
14 kg would have been dissolved.
A variation on this scenario was modeled by placing the
DNAPL on the edge of the capture zone, 60 m from the
warm well, where increased flow velocities were limited. In
such a case, a minor increase of DNAPL dissolution was found
(5.13 g/d) compared with a situation without SATES (1.95 g/d),
but this increase was significantly lower compared with the
standard scenario 3. Since no contaminants entered the well of
the SATES system in this alternative setting, the role of the
SATES system in plume dilution was limited.
3.4. Scenario 4: combined effect of enhanced DNAPL dissolution
and degradation within the SATES system
As expected, the inclusion of degradation in this scenario
4 resulted in a significantly smaller increase in both the
contaminated groundwater volume and dissolved mass of
PCE compared to scenario 3. The DNAPL mass remaining after
19 years in case of SL degradation was slightly less compared
to scenario 3. In the plume, a groundwater volume of
exceeded the TCL for PCE, of which 24.5 Mm
exceeded the MCL. For comparison, if no SATES had been
introduced, this was only 0.7 Mm
(>TCL) and little more
than 0.6 Mm
In Fig. 7A it is shown that the PCE mass increased to 95 kg
in 19 years, whereas 32 kg was reached in a situation
without the SATES system. The rate of PCE mass increase
becomes smaller with time for SL degradation. It is expected
that, ultimately, the mass of PCE would remain approximate-
ly constant in time as long as the DNAPL source remains, with
superimposed seasonal variations caused by varying DNAPL
dissolution in response to alternating recovery and injection
and variable pumping rates. In this situation, the mass of PCE
dissolved each year equals the mass of PCE removed by
degradation. The increased degradation capacity is due to the
increased contaminated aquifer volume. Therefore, the level
of plume mass stabilization depends on the zero-order rate
constant and the spatial extent of the plume.
As with SL degradation, the total mass of PCE stabilized
over time in the CL degradation case (Fig. 7B) as the mass
dissolved from the DNAPL each year approached the mass
removed by degradation. In this case, this can be explained
by the increase in PCE plume mass, which controls the overall
degradation. The level of stabilization depends solely on the
degradation rate constant. Total dissolved PCE mass in the
aquifer increased to 151 kg, but would have decreased to
6 kg if no SATES had been introduced. An aquifer volume of
exceeded the TCL and 103.6 Mm
Fig. 9. Plan view distribution of PCE concentrations exceeding the TCL (6 ×10
mmol/L) and MCL (2.4×10
mmol/L) in scenario 3 at the bottom of aquifer 1.
The cold well is indicated by ‘C’, the warm well by ‘W’. The end of the first cycle at t =1 year (A) and the end of the last cycle (t =19 years, B) is shown.
9K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
MCL for PCE. This would have been 12.3 Mm
(>MCL) if no SATES was introduced.
The results of this study provide important insights for the
evaluation of future contaminant conditions at CHC contam-
inated SATES sites. With the relatively simple descriptions of
the DNAPL dissolution and degradation processes, the
disparate results from the different scenarios clearly illustrate
the interaction between a SATES well pair at a (CHC)
contaminated site on plume concentrations and, total
plume volume. Evidently, specific outcomes such as plume
size and mass development will depend on site-specific
conditions such as rate constants and limiting factors for
degradation, degree of mixing within the SATES system and
the presence and location of DNAPL with respect to the SATES
systems. However, clear generic trends and controlling
processes were identified in this study, of which the most
important ones will be discussed below.
4.1. How degradation kinetics affect contaminant degradation
in SATES systems
In a SL case, a higher degree of solute spreading caused by
the SATES system results in increased degradation marked by
contaminant mass removal, but this beneficial effect is absent
in a case with CL degradation. Furthermore, it is found that
SATES can cause a dramatically expanding PCE plume in
which concentrations exceed the TCL for the CL degradation
case compared to the reference scenario (no SATES), albeit
that the plume volume exceeding MCL is reduced. In contrast,
the plume volumes exceeding both TCL and MCL can
decrease for the SL degradation case, provided that zero-
order rate constants are sufficiently high.
For the SATES configuration and contaminant situation
considered, a dissolvedPCE plume degraded almost completely
in 19 years for the SL case (scenario 2). Therefore, the
degradation rate constants obtained from literature are high
enough for clean-up in case of SL and absence of DNAPL
(scenario 2). In case of DNAPL presence, a steady state situation
was attained for both degradation regimes within decades
(scenario 4, SL and CL).
For the purpose of this study it was assumed that the
redox conditions for full CHC degradation were present
throughout the modeled domain and the SATES system did
not affect those conditions. However, SATES is applied in
aquifers having subtle redox gradients within overall anoxic
conditions (Bonte et al., 2011b). In such aquifers, a stratifi-
cation of degradation rates can be expected as particular
CHCs are more easily degraded in specific redox conditions.
The natural redox stratification may become disturbed with
the introduction of SATES, leading to mixing of methane,
sulfate, and/or nitrate with resultant variability in PCE, TCE,
DCE, and VC degradation (e.g., Bradley, 2000; Bradley and
Chapelle, 2011; Suarez and Rifai, 1999; Wiedemeier et al.,
1999). Although temperature effects on degradation rates
could be ignored for the conditions of thermally balanced
SATES systems with a small temperature difference in this
study, these should be taken into account when studying
thermally non-balanced SATES systems with a considerably
larger temperature difference.
4.2. Contaminant dilution and enhanced degradation
When dissolved CHC plumes are present in a homogeneous
target aquifer and DNAPL is absent, the dilution caused by
mixing in the well is controlled by the proportion of the plume
relative to the aquifer thickness, transmissivity stratification
and the proportion of the capture zone of the well that is
contaminated (Einarson and Mackay, 2001). Consequently,
dilution increases when the ratio of plume thickness over
screen length decreases, or when ambient groundwater flow
velocities are higher, as the reinjected CHC plume will partly
move outside the upstream orientated capture zone of the
well. This is followed by mixing of the recoverable part of the
plume with upstream (uncontaminated) groundwater (Bear
and Jacobs, 1965; Ceric and Haitjema, 2005). As illustrated in
the modeling scenarios, the mixing effect increases overall
degradation when the degradation rate is sediment-limited.
This is due to the SATES system, which reinjects mixed
groundwater with diluted contaminant concentrations and
causes enhanced spreading of contaminants in the aquifer,
exposing the contaminants to a greater portion of degradation
capacity present in the sediment. As a consequence, CHC
concentrations will not only decrease by dilution, but also by
the enhanced mass removal due to spreading by SATES, which
is larger in case of stronger dilution. This study shows how
relevant this coupled dilution and spreading is for the
development of the CHC plume.
4.3. Inﬂuence of DNAPL presence and mass ﬂux
In this study, scenarios with and without a DNAPL source
are evaluated. The equilibrium-controlled DNAPL-dissolution
rate was a function of the difference between its solubility
and the groundwater concentration. Therefore, the degree to
which SATES causes mixing of the contaminant plume with
unpolluted groundwater, and thus maintains low concentra-
tions, will largely control enhanced DNAPL dissolution. This
study shows that when DNAPL is present outside the capture
zone of the SATES system, the influence of the SATES system
on DNAPL dissolution becomes strongly reduced even though
enhanced DNAPL dissolution still occurs due to increased
groundwater flow velocities. Neglecting dispersion and
background lateral flow, mixing of CHC dissolved from
DNAPL by SATES can only occur when the DNAPL is within
a radius r
, which is controlled by the radius of the capture
) and the retardation factor R(Eq. (2)):
This implies that the most mobile component dissolved
from DNAPL or generated by degradation in the plume having
the lowest retardation factor (in this case VC) determines
whether dilution of any CHC species in the SATES wells occurs.
The use of a high lumped rate-transfer coefficient (ω
our simulation of DNAPL dissolution represents the upper-
limit of enhanced mass flux compared to the reference
10 K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
condition with no SATES system. However, due to the
configuration and small size of the DNAPL in our model, the
increased mass discharge (up to 76.6 g/d) due to the presence
of the SATES system can be considered intermediate in
comparison to mass discharge at actual DNAPL contaminated
sites (Newell et al., 2011). The model is highlighting the
worst-case potential relevance of DNAPLs near the SATES
wells for the given DNAPL volume and geometry. In case
dissolution from the DNAPL is mass-transfer limited (Miller et
al., 1990; Park and Parker, 2005) within the SATES affected
DNAPL source zone, the increased groundwater flow velocities
in SATES systems will not result in equally increased mass
fluxes (Parker and Park, 2004).
Depending on the source strength, biodegradation capac-
ity, and the extent to which mass-transfer limited dissolution
occurs, the steady-state contaminant situation found in this
study (Fig. 7A,B) may be established more rapidly, resulting
in a smaller aquifer volume becoming contaminated. How-
ever, as DNAPL dissolution will be slower in time, the CHC
plume will persist longer, albeit at a smaller scale.
4.4. Risk-assessment of SATES as a remediation technology
Performance of different remediation technologies is
evaluated based on multiple methods and standards. Rele-
vant criteria for this study are provided by Rao et al. (2001).
Rao et al. (2001) gauged the performance of remediation by
whether ‘(a) the spatial extent of the existing dissolved
plume was stable or decreased; (b) the total contaminant
mass within the plume was constant or diminishing; (c) both
the average concentration and the range in concentrations
were diminishing; and (d) contaminant fluxes decreased at
successive control planes along the dissolved plume’. In order
to meet condition (b), the contaminant flux leaving the
source zone must be equal to or less than the potential to
degrade the contaminant in the plume.
In our study, the reference scenario 1 failed to meet
condition (a) as the starting plume was not in equilibrium
after 19 years (Table 4). Only the SATES scenario without
DNAPL (scenario 2) met conditions (b), (c), and (d), but
failed to achieve (a) on the short-term. For SL degradation,
condition (a) was met only over the longer term. Therefore,
this study shows that for a dissolved plume with depleted or
removed DNAPL, beneficial effects of SATES are achieved only
for SL degradation and if a temporary plume size increase is
In scenarios 3 and 4, none of the conditions were met in the
short term, although in scenario 4 conditions (a) and (b) were
met eventually. Although the mixing by SATES lowered the
average concentration, the range in concentrations remained,
so condition (c) could not be satisfied. Where DNAPL occurs,
therefore, this study shows that SATES may only be considered
a successful remediation scheme if (1) DNAPL source-zone
depletion is a remediation target and (2) a continuous larger
contaminated aquifer volume is accepted in the long-term.
This study shows the relevance of (information on)
DNAPL presence, and its mass and distribution in the target
aquifer. Further effects of SATES on the contaminant
distribution will depend on the degradation regime (Section
4.1) and is directly related to the hydrogeological setting and
the configuration of the SATES wells relative to (potential)
DNAPL zones (Section 4.2). This study also highlights the
importance of long-term monitoring during field application
in order to evaluate its impacts, as an increased groundwater
contamination may precede ultimate plume attenuation.
This is the first comprehensive study of the interaction
between SATES pumping and CHC in a groundwater system.
Using a reactive transport modeling approach, conditions that
either promote or negate remediating effects of SATES were
demonstrated. Only if the degradation of contaminants is
controlled by sediment reactivity, the spreading of contami-
nants strongly enhances contaminant mass removal compared
with a situation without SATES, provided that (redox)
conditions for degradation do not deteriorate due to the
SATES system. However, in case degradation is controlled by
the contaminant concentration, a larger plume is generated
without any enhanced mass removal. This study also shows
that under conditions of partially increased DNAPL dissolution,
particularly near SATES wells, a potentially long-lasting
increase in both plume volume and mass may be anticipated.
The degree to which these SATES-influenced plumes reach
spatially stable conditions will depend strongly on the mass
discharge from the DNAPL source, in situ geochemistry, and
degradation rates in the aquifer. Overall, the interaction
between the design and operational conditions of SATES
Risk-assessment SATES in contaminated aquifers according to Rao et al. (2001) for the sediment-limited (SL) and contaminant-limited (CL) degradation cases.
Scenario Degradation Condition a Condition b Condition c Condition d
Spatial extent Contaminant mass Average and range of concentration Contaminant fluxes
S-1 SL −+ + +
CL −+ + +
S-2 SL +/−+ + +
CL −+ + +
S-3 –− − − −
S-4 SL +/−+/− − −
CL +/−+/− − −
+: condition is met.
+/−: condition is met only on the long term.
-: condition is not met.
11K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
systems and site-specific aspects of hydrogeological, biogeo-
chemical, and contaminant characteristics within an aquifer,
determines the quantitative effect of SATES systems on
contaminant degradation and spreading. In the assessment of
further potential benefits or risks of SATES in CHC contami-
nated aquifers, insight into the occurrence of biodegradation
under varying temperatures and mixed redox zones, as well as
the presence and location of DNAPL within the aquifer is
This study was performed within the Dutch national
research project ‘Meer met Bodemenergie’(MMB). The editor
and four anonymous reviewers are thanked for their thought-
ful comments that helped to improve this manuscript.
Arrhenius, S., 1889. Uber die Reaktiongeschwindigkeit bei der Inversion von
Rohrzucker durch sauren. Zeitschrift fur Physik Chemique 4, 226–248.
Bear, J., Jacobs, M., 1965. On the movement of water bodies injected into
aquifers. Journal of Hydrology 3 (1), 37–57.
Bekins, B.A., Warren, E., Godsy, E.M., 1998. A comparison of zero-order, first-
order, and monod biotransformation models. Ground Water 36 (2),
Bonte, M., Stuyfzand, P.J., Van den Berg, G.A., Hijnen, W.A.M., 2011a. Effects
of aquifer thermal energy storage on groundwater quality and the
consequences for drinking water production: a case study from the
Netherlands. Water Science and Technology 63, 1922–1931.
Bonte, M., Stuyfzand, P.J., Hulsman, A., Van Beelen, P., 2011b. Underground
thermal energy storage: environmental risks and policy developments
in the Netherlands and EU. Ecology and Society 16 (1), 22 (online).
Bradley, P.M., 2000. Microbial degradation of chloroethenes in groundwater
systems. Hydrogeology Journal 8 (1), 104–111.
Bradley, P.M., Chapelle, F.H., 2011. Microbial mineralization of dichloroethene
and vinyl chloride under hypoxic conditions. Ground Water Monitoring &
Remediation 31 (4), 39–49.
Brons, H.J., Griffioen, J., Appelo, C.A.J., Zehnder, A.J.B., 1991. (Bio)geochemical
reactions in aquifer material from a thermal energy storage site. Water
Research 25 (6), 729–736.
Carotenuto, A., Fucci, F., La Fianza, G., Reale, F., 1991. Physical model and
demonstration of an aquifer thermal energy store. Heat Recovery
Systems and CHP 11 (2–3), 169–180.
Ceric, A., Haitjema, H., 2005. On using simple time-of-travel capture zone
delineation methods. Ground Water 43 (3), 408–412.
Dillon, P., Pavelic, P., Toze, S., Rinck-Pfeiffer, S., Martin, R., Knapton, A.,
Pidsley, D., 2006. Role of aquifer storage in water reuse. Desalination 188
Edworthy, K.J., Puri, S., 1986. Groundwater and aquifers; an overview of
‘exotic’uses. Quarterly Journal of Engineering Geology & Hydrogeology
19 (2), 87–95.
Einarson, M.D., Mackay, D.M., 2001. Predicting impacts of groundwater
contamination. Environmental Science & Technology 35 (3), 66A–73A.
European Environment Agency, 2007. State of the Environment. EEA,
Evans, D., Stephenson, M., Shaw, R., 2009. The present and future use of
‘land’below ground. Land Use Policy 26 (Supplement 1), S302–S316.
Faneca-Sànchez, M., Hermeler, H., Huibregtse, E., Mayer, M., Meindertsma,
W., Van Oort, S., Sommer, W., Sun, W., Vermeer, M., Hao, W., 2010. A
Method for Sustainable Water Management —Test Case ‘De Uithof’
TNO-034-UT-2010-01059. TNO, Utrecht.
Ferguson, B.K., 1990. Urban storm water infiltration. Journal of Soil and
Water Conservation 45 (6), 605–609.
Foster, S.S.D., 2001. The interdependence of groundwater and urbanisation
in rapidly developing cities. Urban Water 3 (3), 185–192.
Friis, A.K., Heimann, A.C., Jakobsen, R., Albrechtsen, H.-J., Cox, E., Bjerg, P.L.,
2007. Temperature dependence of anaerobic TCE-dechlorination in a
highly enriched Dehalococcoides-containing culture. Water Research 41
Gao, Q., Li, M., Yu, M., Spitler, J.D., Yan, Y.Y., 2009. Review of development
from GSHP to UTES in China and other countries. Renewable and
Sustainable Energy Reviews 13 (6–7), 1383–1394.
Halford, K.J., Hanson, R.T., Santa Clara Valley Water, D., 2002. User Guide for
the Drawdown-Limited, Multi-node Well (MNW) Package for the US
GeologicalSurvey's Modular Three-dimensional Finite-difference Ground-
water Flow Model,Versions MODFLOW-96and MODFLOW-2000.US Dept.
of the Interior, US Geological Survey.
Harbough, A.W., Banta, E.R., Hill, M.C., McDonald, M.G., 2000. Modflow-2000,
the U.S. Geological Survey Modular Groundwater Model —User Guide to
Modularization Concepts and the Groundwater Flow Process. Open-File
Report 00–92. U.S. Geological Survey.
Hartog, N., 2011. Anticipated temperature effects on biogeochemical
reaction rates in seasonal aquifer thermal energy storage (ATES)
systems: an evaluation using the Arrhenius equation. 1ste Nationaal
Congres Bodemenergie,Maarssen (http://www.nielshartog.nl/articles/
Hartog, N., Grif fioen, J. , van Bergen, P.F., 2005. Deposi tional an d
paleohydrogeological controls on the distribution of organic matter and
other reactive reductants in aquifer sediments. Chemical Geology 216
Hayden, N.J., Voice, T.C., Annable, M.D., Wallace, R.B., 1992. Prediction of
leachate concentrations in petroleum-contaminated soils. Journal of Soil
Contamination 1 (1), 81–93.
Jakobsen, R., Postma, D., 1994. In situ rates of sulfate reduction in an aquifer
(Rømø, Denmark) and implications for the reactivity of organic matter.
Geology 22 (12), 1101–1106.
Karickhoff, S.W., 1981. Semi-empirical estimation of sorption of hydrophobic
pollutants on natural sediments and soils. Chemosphere 10, 833–846.
Kim, J., Lee, Y., Yoon, W.S., Jeon, J.S., Koo, M.-H., Keehm, Y., 2010. Numerical
modeling of aquifer thermal energy storage system. Energy 35 (12),
KNMI, 2011. Weather database: Station De Bilt. KNMI, De Bilt.
Langevin, C.D., Thorne, D.T., Dausman, A.M., Sukop, M.C., Guo, W., 2008b.
SEAWAT version 4: a computer program for simulation of multi-species
solute and heat transport. in: U.S.G. Survey (Editor), Reston, Virginia.
Leonard, B.P., 1988. Universal limiter for transient interpolation modeling of
the advective transport equations: the ULTIMATE conservative differ-
ence scheme. NASA Technical Memorandum 100916 ICOMP-88-11.
Ma, R., Zheng, C., 2010. Effects of density and viscosity in modeling heat as a
groundwater tracer. Ground Water 48 (3), 380–389.
Mao, X., Prommer, H., Barry, D.A., Langevin, C.D., Panteleit, B., Li, L., 2006.
Three-dimensional model for multi-component reactive transport with
variable density groundwater flow. Environmental Modelling & Soft-
ware 21 (5), 615–628.
McCarty, P.L., 1997. Breathing with chlorinated solvents. Science 276 (5318),
Miller, C.T., Poirier-McNeil, M.M., Mayer, A.S., 1990. Dissolution of trapped
nonaqueous phase liquids: mass transfer characteristics. Water Re-
sources Research 26 (11), 2783–2796.
Molz, F.J., Warman, J.C., Jones, T.E., 1978. Aquifer storage of heated water:
part I—a field experiment. Ground Water 16 (4), 234–241.
Molz, F.J., Parr, A.D., Andersen, P.F., Lucido, V.D., Warman, J.C., 1979. Thermal
energy storage in a confined aquifer: experimental results. Water
Resources Research 15 (6), 1509–1514.
Molz, F.J., Parr, A.D., Andersen, P.F., 1981. Thermal energy storage in a confined
aquifer: second cycle. Water Resources Research 17 (3), 641–645.
Molz, F.J., Melville, J.G., Parr, A.D., King, D.A., Hopf, M.T., 1983. Aquifer
thermal energy storage: a well doublet experiment at increased
temperatures. Water Resources Research 19 (1), 149–160.
Montgomery, J.H., 2000. Groundwater Chemicals —Desk Reference. CRC
Press, Boca Raton, Florida, USA.
Newell, C.J., Farhat, S.K., Adamson, D.T., Looney, B.B., 2011. Contaminant
plume classification system based on mass discharge. Ground Water 49
Park, E., Parker, J.C., 2005. Evaluation of an upscaled model for DNAPL
dissolution kinetics in heterogeneous aquifers. Advances in Water
Resources 28 (12), 1280–1291.
Parker, J.C., Park, E., 2004. Modeling field-scale dense nonaqueous phase
liquid dissolution kinetics in heterogeneous aquifers. Water Resources
Research 40 (5), W05109.
Parker, B.L., Cherry, J.A., Chapman, S.W., Guilbeault, M.A., 2003. Review and
analysis of chlorinated solvent dense nonaqueous phase liquid distribu-
tions in five sandy aquifers. Vadose Zone Journal 2 (2), 116–137.
Parkhurst, D.L., Appelo, C.A.J., 1999. User's guide to PHREEQC (version 2): a
computer program for speciation, batch-reaction, one-dimensional
transport, and inverse geochemical calculations. Water-Resources In-
vestigations Report. : Open-File Reports Section [Distributor]. U.S.
Geological Survey: Earth Science Information Center, Denver, Colorado,
Pavlostathis, S.G., Zhuang, P., 1993. Reductive dechlorination of chloroalkenes
in microcosms developed with a field contaminated soil. Chemosphere 27
12 K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13
Post, V.E.A., Prommer, H., 2007. Multicomponent reactive transport simula-
tion of the Elder problem: effects of chemical reactions on salt plume
development. Water Resources Research 43 (10), W10404.
Prommer, H., Post, V.E.A., 2010. A Reactive Multicomponent Model for
Saturated Porous Media, Version 2.0, User's Manual. http://www.pht3d.
Prommer, H., Davis, G.B., Barry, D.A., 1999. Geochemical changes during
biodegradation of petroleum hydrocarbons: field investigations and
biogeochemical modelling. Organic Geochemistry 30 (6), 423–435.
Prommer, H., Barry, D.A., Zheng, C., 2003. MODFLOW/MT3DMS-based
reactive multicomponent transport modeling. Ground Water 41 (2),
Pyne, R.D.G., 2005. Aquifer Storage and Recovery —A Guide to Groundwater
Recharge Through Wells. ASR Systems LLC, Gainesville, Florida (608
Rao, P.S.C., Jawitz, J.W., Enfield, G.E., Falta, R.W., Annable, M.D., Wood, A.L.,
2001. Technology integration for contaminated site remediation: clean-
up goals and performance criteria. Groundwater Quality.Sheffield, U.K.
Robinson, C., Brovelli, A., Barry, D.A., Li, L., 2009. Tidal influence on BTEX
biodegradation in sandy coastal aquifers. Advances in Water Resources
32 (1), 16–28.
Sanner, B., Karytsas, C., Mendrinos, D., Rybach, L., 2003. Current status of
ground source heat pumps and underground thermal energy storage in
Europe. Geothermics 32 (4–6), 579–588.
Slenders, H.L.A., Dols, P., Verburg, R., de Vries, A.J., 2010. Sustainable
remediation panel: sustainable synergies for the subsurface: combining
groundwater energy with remediation. Remediation Journal 20 (2),
Smidt, H., de Vos, W.M., 2004. Anaerobic microbial dehalogenation. Annual
Review of Microbiology 58 (1), 43–73.
Steeneveldt, R., Berger, B., Torp, T.A., 2006. CO2 capture and storage: closing
the knowing–doing gap. Chemical Engineering Research and Design 84
Suarez, M.P., Rifai, H.S., 1999. Biodegradation rates for fuel hydrocarbons and
chlorinated solvents in groundwater. Bioremediation Journal 3 (4),
Taskforce ATES, 2009. Green Light for Underground Thermal Energy Storage.
(in Dutch) Ministry of Housing, Spatial Planning and Envrironment
(VROM), Den Haag (downloadable from: www.vrom.nl).
TNO-NITG, 2009. REGIS II.1 —Regional Hydrogeological Model. www.
Tsang, C.F., Hopkins, D.L., 1982. Aquifier Thermal Energy Storage: A Survey
(Medium: X; Size: Pages: 427–441 pp.).
Tsang, C.F., Birkholzer, J., Rutqvist, J., 2008. A comparative review of
hydrologic issues involved in geologic storage of CO
disposal of liquid waste. Environmental Geology 54 (8), 1723–1737.
Van Breukelen, B.M., Hunkeler, D., Volkering, F., 2005. Quantification of
sequential chlorinated ethene degradation by use of a reactive transport
model incorporating isotope fractionation. Environmental Science &
Technology 39 (11), 4189–4197.
van Helvoort, P.-J., Griffioen, J., Hartog, N., 2007. Characterization of the
reactivity of riverine heterogeneous sediments using a facies-based
approach; the Rhine–Meuse delta. (The Netherlands) Applied Geo-
chemistry 22 (12), 2735–2757.
Verburg, R., Slenders, H.L.A., Hoekstra, N., Van Nieuwkerk, E., Guijt, R., Van
der Mark, B., 2010. Manual BOEG: Underground Thermal Energy and
Groundwater Contamination (in Dutch).
VROM, 2009. Remediation Circular 2009. in: Ministry of Housing, Spatial
Planning and Environment (VROM) (Editor), BWBR0025649, The
Wiedemeier, T.H., Rifai, H.S., Wilson, J.T., Newell, C.J., 1999. Natural
attenuation of fuels and chlorinated solvents in the subsurface. John
Wiley and Sons, New York.
Zheng, C., 2006. MT3DMS v5. 2 Supplemental User's Guide. Department of
Geological Sciences, University of Alabama, Tuscaloosa, AL.
Zheng, C., Wang, P.P., 1999. MT3DMS: a modular three-dimensional
multispecies transport model for simulation of advection, dispersion
and chemical reactions of contaminants in groundwater systems;
Documentation and user's guide. in: U.S. Army Corps of Engineers
(Editor), Tuscaloosa, Alabama.
13K.G. Zuurbier et al. / Journal of Contaminant Hydrology 147 (2013) 1–13