monitoring trichloroethene remediation at an iron permeable reactive barrier using stable carbon...

Journal of Contaminant Hydrology 78 (2005) 313–325

www.elsevier.com/locate/jconhyd

Monitoring trichloroethene remediation at an iron

permeable reactive barrier using stable carbon

isotopic analysis

Nancy VanStone a,b, Andrzej Przepiora c, John Vogan c,

Georges Lacrampe-Couloume a, Brian Powers d, Ernesto Perez e,

Scott Mabury b, Barbara Sherwood Lollar a,*

aStable Isotope Laboratory, University of Toronto, 22 Russell St., Toronto, ON, Canada, M5S 3B1bDepartment of Chemistry, University of Toronto, 80 St. George St., Toronto, ON, Canada, M5S 3H6

cEnviroMetal Technologies, Inc., 745 Bridge St. W., Suite 7, Waterloo, ON, Canada, N2V 2G6dURS Corporation, 8181 East Tufts Ave., Denver, CO, 80237, United States

eAFCEE/ ERD-FEW, 300 Vesle Drive, Building 367, F.E. Warren AFB, WY, 82005, United States

Received 27 May 2004; received in revised form 31 May 2005; accepted 31 May 2005

Abstract

Stable carbon isotopic analysis, in combination with compositional analysis, was used to evaluate

the performance of an iron permeable reactive barrier (PRB) for the remediation of ground water

contaminated with trichloroethene (TCE) at Spill Site 7 (SS7), F.E. Warren Air Force Base,

Wyoming. Compositional data indicated that although the PRB appeared to be reducing TCE to

concentrations below treatment goals within and immediately downgradient of the PRB,

concentrations remained higher than expected at wells further downgradient (i.e. N9 m) of the

PRB. At two wells downgradient of the PRB, TCE concentrations were comparable to upgradient

values, and d13C values of TCE at these wells were not significantly different than upgradient values.

Since the process of sorption/desorption does not significantly fractionate carbon isotope values, this

suggests that the TCE observed at these wells is desorbing from local aquifer materials and was

present before the PRB was installed. In contrast, three other downgradient wells show significantly

more enriched d13C values compared to the upgradient mean. In addition, d13C values for the

degradation products of TCE, cis-dichloroethene and vinyl chloride, show fractionation patterns

0169-7722/$ -

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ding author. Tel.: +1 416 978 0770; fax: +1 416 978 3938.

ress: [email protected] (B. Sherwood Lollar).

N. VanStone et al. / Journal of Contaminant Hydrology 78 (2005) 313–325314

expected for the products of the reductive dechlorination of TCE. Since concentrations of both TCE

and degradation products drop to below detection limit in wells within the PRB and directly below it,

these downgradient chlorinated hydrocarbon concentrations are attributed to desorption from local

aquifer material. The carbon isotope values indicate that this dissolved contaminant is subject to local

degradation, likely due to in situ microbial activity.

D 2005 Elsevier B.V. All rights reserved.

Keywords: Carbon isotopes; Groundwater remediation; Permeable reactive barrier; Reductive dechlorination;

Trichloroethene; Zero-valent iron

1. Introduction

Permeable reactive barriers (PRBs) constructed of elemental iron have emerged as an

effective passive remediation method for ground water contaminated with a range of

contaminants, mainly chlorinated hydrocarbons (CHCs) (e.g. Agrawal and Tratnyek,

1996; Alowitz and Scherer, 2002; Gillham and O’Hannesin, 1994; Hozalski et al., 2001;

Nam and Tratnyek, 2000). A PRB is an in situ engineered zone of reactive material placed

across the path of contaminated ground water. The major advance of PRBs over other

ground water remediation approaches is the lack of above ground structures, low operation

and maintenance cost and the enhanced remediation efficiency, particularly compared with

pump-and-treat systems. There are currently more than 90 PRBs installed at CHC

contaminated sites (O’Hannesin, 2003). Many studies have investigated the pathways,

mechanisms, kinetics and longevity of CHC degradation using elemental iron (see Scherer

et al., 2000; Tratnyek, 2002).

Assessment of iron PRB performance is based on monitoring CHC concentrations,

along with pH and Eh, and major inorganic constituents, in ground water well transects

across the PRB. Often, iron PRBs are installed within existing contaminant plumes and

therefore elevated concentrations of contaminants are observed downgradient of PRBs for

some time after the system has been installed, depending on the extent of initial

contamination, ground water flow rates, desorption rates and type of the aquifer material.

The persistence of CHCs downgradient of iron PRBs is expected because significant

volumes of ground water may be needed for the remnant contaminants to be desorbed and

flushed from downgradient aquifer materials (Heneman et al., 2001; Powell et al., 1998). It

is difficult, however, to evaluate PRB performance in situations where elevated

contaminant concentrations downgradient of the PRB remain for longer than expected

periods (Powell et al., 1998). This could arise for three reasons: (1) incomplete degradation

within the PRB, (2) hydraulic bypass underneath or around the PRB, and (3) slower than

expected desorption of remnant contaminants trapped within aquifer materials. Distin-

guishing between these processes is important for the assessment of PRB performance and

consequently developing new techniques that help resolve these processes is of

considerable interest.

Fractionation of stable carbon isotopes has been observed for chlorinated ethenes

during processes such as biodegradation (Bloom et al., 2000; Hunkeler et al., 1999;

Sherwood Lollar et al., 2001; Slater et al., 2001), oxidation by permanganate (Hunkeler et

N. VanStone et al. / Journal of Contaminant Hydrology 78 (2005) 313–325 315

al., 2003; Poulson and Naraoka, 2002), and reductive dechlorination on zero-valent iron

(Bill et al., 2001; Dayan et al., 1999; Slater et al., 2002; VanStone et al., 2004). Carbon

isotopic fractionation results from differences in the rates of reaction for 13C- and 12C-

bearing molecules. The differences in reaction rates for the different isotopes are due to

mass-dependent differences in activation energies for the respective reactions (Fry, 1971).

In general, for most chlorinated ethenes and aromatic compounds, the lighter isotope (12C)

reacts faster than the heavier isotope (13C), leading to fractionation, and enrichment of the

heavy isotope in the remaining reactant as the reaction proceeds. Likewise, this leads to

enrichment in the light isotope with respect to the parent compound in the products of the

reaction. Laboratory experiments have shown that carbon isotopic fractionation for

chlorinated ethenes is not significant (i.e. b0.5x) for non-degradative processes such as

dissolution, vaporization or adsorption under equilibrium conditions (Dempster et al.,

1997; Harrington et al., 1999; Poulson and Drever, 1999; Slater et al., 1999, 2000),

whereas carbon isotopic fractionation associated with degradative processes is on the order

of tens of permil (x).

Carbon compound specific isotopic analysis (CSIA) is used to measure carbon isotope

fractionation and has been shown to be a powerful tool for identifying in situ

biodegradation in the field for contaminants such as perchloroethene (PCE) and

trichloroethene (TCE) (Hunkeler et al., 1999; Sherwood Lollar et al., 2001; Song et al.,

2002; Vieth et al., 2003). Several studies have shown large and reproducible carbon

isotopic fractionation during the degradation of chlorinated ethenes on electrolytic and cast

iron (Bill et al., 2001; Dayan et al., 1999; Slater et al., 2002; VanStone et al., 2004), but

these observations have not been applied in the field to date. The objective of this study

was to investigate the use of carbon stable isotopic analysis to help evaluate PRB

performance at a field installation for the remediation of TCE. Routine performance

monitoring of the PRB consisted of quantification of TCE and its products upgradient,

within and downgradient of the iron PRB. In addition, carbon CSIAwas carried out in the

vicinity of the PRB to investigate the source of elevated chlorinated ethene concentrations

observed downgradient of the installation.

2. Site background

A PRB constructed of iron filings was installed in 1999 at the F.E. Warren Air Force

Base, Wyoming, Spill Site 7 (SS7) to remediate ground water contaminated with TCE.

The TCE originates from a defunct on-site liquid oxygen manufacturing facility. A grease

trap at the plant was the source of organic solvents (primarily TCE) to a surface drainage

ditch, resulting in migration of the solvents through the vadose zone to local ground water

(Heneman et al., 2001). Construction of the PRB was completed in October 1999. A

schematic of the SS7 PRB is shown in Fig. 1, with the direction of ground water flow

indicated. The PRB is 173 m long, has a flow through thickness of 1.2 m and a saturated

depth of 4.6 m below the historic low ground water level, with a low-permeability clay cap

installed on the PRB to minimize ground water flow over the top of the treatment PRB.

The PRB was installed at the site as an interim remediation measure and it does not extend

to the bottom of the contaminated shallow aquifer. The PRB is located within 20 to 60 m

Fig. 1. Schematic Map of SS7 PRB. Gray-filled circles indicate the position of monitoring wells that contained

concentrations of contaminants above the detection limits (d.l.) for isotopic analysis, and black-filled circles

indicate wells that did not contain adequate concentrations for CSIA. The treatment PRB is approximately 173 m

long. Base map is to scale (1 cm=10 m). Hatched lines show expanded view of transects 1, 2 and 3. Section 1 is

1.2 m thick, Section 2 is 0.3 m thick and Section 3 is 0.5 m thick. Expanded views show sampling wells located at

the centre of each main section (Transects 1 to 3). See text for detailed description of each monitoring well. Wells

indicated (*) are at a depth at least 1 m below the PRB.


of a perennial creek. The SS7 PRB contains approximately 1.6�106 kg of iron filings. To

accommodate the variations in ground water velocity and CHC concentration expected

along the line of PRB installation, the amount of iron in the PRB was varied by using a

different proportion of iron/sand in each of the three sections: Section 1 with 100% iron,

Section 2 with 25% iron/75% sand mix, and Section 3 with 38% iron/62% sand mix.

After the PRB was installed in October 1999, the system was left undisturbed for 6

months to allow the ground water conditions to equilibrate before implementing a

performance monitoring program (Heneman et al., 2001). The monitoring wells discussed

in this study are shown in Fig. 1 and represent only a portion of the wells at the site. These

monitoring wells are made of PVC and have 1.5 m screens. There are a series of

monitoring wells at the centre of each segment of the treatment PRB (Transects 1, 2 and

3), which are shown in expanded view in Fig. 1. In Transect 1, wells 101A and 101B are

located within a meter of the upgradient side of the PRB at depths of 8.5 and 4.5 m,

respectively, with well 101A located about 1 m below the PRB and 101B located within


the ground water plume transected by the PRB. Well 102 is located in the middle of the

PRB. Wells 103A and 103B are located approximately 1 m downgradient of the PRB at

depths of 8.5 and 4.5 m, respectively, with 103A located about 1 m below the PRB and

103B located within the ground water plume intercepted by the PRB. Well 700B is located

9 m downgradient from the PRB. The wells for the other 2 transects (Transect 2 and 3) are

similarly located with respect to the PRB.

Periodic sampling of the groundwater monitoring wells has been conducted since April

2000 for concentrations of CHCs (TCE, cDCE, trans-dichloroethene and VC) using

standard EPA method 502.1, and several inorganic indicators of water quality (Heneman et

al., 2001). There were consistent decreases in Eh, sulfate, calcium and magnesium

concentrations and increases in pH consistent with those documented at other PRB

projects (O’Hannesin and Gillham, 1998).

Performance monitoring of the SS7 PRB revealed a decrease in CHC concentration

across the PRB from upgradient to non-detectable values within the PRB and immediately

downgradient (Heneman et al., 2001). In contrast, in wells further downgradient of the

PRB (e.g. wells 186, 700B, 707B and 708 located 9 to 12 m from the PRB) concentrations

of both TCE and cDCE were comparable to upgradient concentrations and were suggested

to be due to desorption of TCE from aquifer materials downgradient of the PRB (Heneman

et al., 2001). Corresponding increases in pH and decreases in Eh, carbonate alkalinity,

Ca2+, and Mg2+ were noted within the PRB and immediately downgradient in comparison

to upgradient wells (Heneman et al., 2001). These trends are expected due to reducing

conditions created by iron corrosion and are consistent with those noted at other PRB sites

(O’Hannesin and Gillham, 1998). The values of iron-sensitive parameters (i.e. Eh and pH)

and constituents (Ca2+, alkalinity) measured in wells 700B, 707B, and 173B were similar

to the upgradient values. It was unclear whether the persistent high concentrations of TCE

and cDCE in the downgradient wells were due to incomplete degradation of TCE within

the PRB, or to continued desorption from the aquifer material (Heneman et al., 2001).

Based on hydraulic heads measured at each monitoring well, ground water flow paths in

the vicinity of the PRB were not significantly altered from pre-PRB ground water flow

records (Heneman et al., 2001), ruling out hydraulic bypass of the PRB as the source of

elevated downgradient concentrations of CHCs.

3. Methods

Samples for concentrations of chlorinated hydrocarbons and carbon isotopic analysis

were taken in June 2002 (Table 1). Four 40 mL samples from each well were preserved

with several drops of undiluted HCl to inhibit microbial activity, and stored without

headspace. Additional water samples were taken from each well for analysis of TCE,

cDCE and VC concentrations by EPA method 502.1. The detection limit for this analysis

is 0.05 ug/L. All samples for carbon CSIA were packed on ice and shipped to the Stable

Isotope Laboratory at the University of Toronto, Canada, and analyzed within 1 month of

sampling. Detection limits for isotopic analysis varies from compound to compound but is

between 10 and 30 ug/L for TCE, cDCE and VC. One day prior to analysis, 2 of each 40

mL water samples were transferred to 160 mL glass bottles with 40 g of NaCl to facilitate

Table 1

Concentrations and stable carbon isotopic compositions of VC, cDCE and TCE from monitoring wells at

F.E.Warren AFB spill site 7 iron-filings PRB

Well TCE cDCE VC

(Ag/L) d13C (x) (Ag/L) d13C (x) (Ag/L) d13C (x)

Upgradient wells

151 14000 �24.4 260 �23.6 bd.l. bd.l.

182 1100 �23.8 95 �20.2 bd.l. bd.l.

190 820 �24.0 19 �23.0 bd.l. bd.l.

101A 1.8 bd.l. bd.l. bd.l. bd.l. bd.l.

101B 1600 �23.8 110 �24.0 bd.l. bd.l.

201A 290 �25.1 6.8 bd.l. bd.l. bd.l.

201B 4600 �23.9 320 �23.2 bd.l. bd.l.

301A 3.7 bd.l. 7.7 bd.l. bd.l. bd.l.

301B 920 �24.1 12 bd.l. bd.l. bd.l.

Average �24.2 �22.8

SD 0.5 1.5

Wells lateral to PRB

401 78 �25.2 6.5 bd.l. bd.l. bd.l.

501 890 �22.6 270 �25.88 9.2 bd.l.

Wells within PRB, and downgradient wells

102 b d.l. bd.l. bd.l. bd.l. bd.l. bd.l.

103A bd.l. bd.l. bd.l. bd.l. bd.l. bd.l.

103B bd.l. bd.l. bd.l. bd.l. bd.l. bd.l.

202 0.2 bd.l. 2.2 bd.l. bd.l. bd.l.

203A 270 bd.l. 19 bd.l. bd.l. bd.l.


302 bd.l. bd.l. bd.l. bd.l. bd.l. bd.l.

303A 0.2 bd.l. 0.3 bd.l. bd.l. bd.l.


173B 25 �16.8 22 �34.8 1.3 bd.l.

186 31 �7.0 120 �9.6 62 �18.5

700B 980 �21.8 200 �29.0 bd.l. bd.l.

707B 3200 �24.1 330 �23.3 bd.l. bd.l.

708 890 �23.7 73 �24.9 bd.l. bd.l.


favourable partitioning of the chlorinated ethenes to the headspace as described by Slater

et al. (1999). The bottles were sealed with PTFE-lined gray butyl septa and aluminum

crimp seals, hand shaken for one minute, and left for 24 h to equilibrate the chlorinated

ethenes into the headspace of the bottles for CSIA.

At the time of analysis, chlorinated ethenes from the headspace above the aqueous

solution was continually extracted from solution by flushing helium (minimum flow rate

of 1.5 mL/min) through the headspace within each sample bottle via 2 extraction needles

piercing the PTFE septa after the method of Morrill et al. (2004). The chlorinated ethenes

were cryogenically trapped in a fused silica loop (34 cm in length, 0.53 mm i.d.) by

immersion in a liquid nitrogen bath (�196 8C). After 8 min of extraction, the 1 m loop

was removed from the liquid nitrogen and placed in boiling water, releasing the


chlorinated ethenes onto the GC column in a single pulse. A DB-624 capillary column (30

m�0.25 mm i.d., 1.4 um film thickness) was used to separate chlorinated ethenes with a

temperature program of 40 8C for 4 min followed by an increase to 90 8C at a rate of 15

8C/min and a final holding time of 7 min and a flow rate of 1.2 mL/min. Injection

temperature was 185 8C. Samples were prepared and analyzed in duplicate to ensure

reproducibility, which was always less than F0.4x. Total error associated with this

method is F0.5x, and incorporates both accuracy and reproducibility (Dempster et al.,

1997; Mancini et al., 2002; Morrill et al., 2004). Carbon isotopic compositions are

reported as d13C, where d13C=(Rs�Rst) /Rst�1000, and Rs and Rst are the ratios of13C

to 12C in the sample and a standard (calibrated to V-PDB), respectively (Clark and Fritz,

1997). d13C is reported in units of permil (x).

4. Results and discussion

Concentrations in upgradient wells (Table 1) range from 290 to 14000 Ag/L for TCE,

and 12 to 320 Ag/L for cDCE. cDCE may have been present in the initial TCE spill as a

trace co-contaminant. Alternatively, cDCE may be present as a product of TCE

degradation by reductive dechlorination. The carbon isotopic composition of TCE in

the upgradient wells is quite uniform, from �23.8x to �25.1x (mean: �24.2F0.5x,

n =7). Well 401, located several metres to the side of the PRB is also similar (�25.2x).

These values are slightly enriched in 13C compared to results obtained for TCE from

industrial sources. The d13C of free product TCE ranges from �31.9x to �25.8x(Beneteau et al., 1999; Shouakar-Stash et al., 2003; van Warmerdam et al., 1995). Physical

processes such as dissolution, dispersion, vaporization or adsorption are not known to

cause significant fractionation for chlorinated ethenes (Dempster et al., 1997; Harrington

et al., 1999; Slater et al., 1999, 2000). Hence, both these enriched d13C values, and the

presence of cDCE suggests TCE upgradient and away from the influence of the PRB may

already be somewhat degraded with respect to the source TCE. A second lateral well

(501), which is located several meters to the side of the PRB, in fact, shows significant

levels of VC (9.2 Ag/L). The occurrence of VC in well 501 may be due to local microbial

activity, as VC is a product of reductive dechlorination by certain microbial communities

(Wiedemeier et al., 1999). TCE from well 501 has a more enriched value of �22.6x with

respect to the mean upgradient value of �24.2x, which, combined with the occurrence of

VC in this well (a product of microbial degradation of TCE) provides further evidence of

some degree of microbial activity in the vicinity of this well, and potentially throughout

the site.

Downgradient of the PRB, a range of TCE concentrations are observed from 25 to 3200

Ag/L (Table 1). The concentrations of TCE measured in wells 700B (Fig. 2a), 707B (Fig.

2b) and 708 (Table 1) are not substantially lower than upgradient of the PRB. In addition,

the carbon isotopic compositions of TCE in wells 707B (�24.1x) and 708 (�23.7x) are

not significantly different than those measured upgradient of the PRB (upgradient mean is

�24.2x) (Table 1). It is clear from Fig. 2b, however, that this is not due to incomplete

degradation within the PRB, as TCE concentrations within the PRB (wells 202 and 203B)

and immediately downgradient are below detection limits. This same trend is true for

Fig. 2. Concentration of cDCE and TCE for Transects 1, 2 and 3 of SS7 PRB. For each transect of the PRB as

shown in Fig. 1, concentrations in Ag/L and isotopic compositions of cDCE and TCE (where applicable) at each

well are shown. Identity of well is indicated at top of horizontal axis. Total error is F5 % for concentration, and

F0.5x for d13C values, incorporating both accuracy and reproducibility after Mancini et al. (2002) and Gray et

al. (2002).


cDCE, which also shows identical d13C values upgradient and downgradient of the PRB,

although concentrations drop to below detection limits within the PRB (Table 1, Fig. 2b).

There is considerable research ongoing into the persistence of CHCs downgradient of

PRBs, including the examination of phenomena such as slow desorption of CHCs from

aquifer sediments and incomplete flushing of aquifer materials due to aquifer


heterogeneities. Assuming that the TCE sorbed to aquifer materials downgradient is from

the same source as the upgradient material, they should have similar carbon isotopic

compositions (Fig. 2b) as partitioning does not show significant fractionation (i.e.

enrichment or depletion of 13C in contaminant pool) for TCE (Beneteau et al., 1999;

Dempster et al., 1997; Poulson and Drever, 1999; Slater et al., 1999, 2000). In contrast,

significant enrichment of 13C in residual TCE is observed during degradative reactions,

such as biologically mediated reductive dechlorination (Bloom et al., 2000; Hunkeler et

al., 1999; Sherwood Lollar et al., 1999; Slater et al., 2001), oxidation by permanganate

(Hunkeler et al., 2003; Poulson and Naraoka, 2002) or abiotic degradation of TCE by iron

(Bill et al., 2001; Dayan et al., 1999; Slater et al., 2002; VanStone et al., 2004). The

similarity between the carbon isotopic values of the downgradient TCE (�24.1x and

�23.7x for wells 707B and 708) and the mean upgradient value, therefore indicates that

the TCE observed in these downgradient wells has not undergone any degradative

processes compared to the TCE observed upgradient. The source of TCE observed in wells

707B and 708, therefore, is likely desorption from the downgradient aquifer material,

indicating that substantial amounts of TCE were present before the PRB was installed, and

it is continuing to desorb from the local materials. An alternative hypothesis for the source

of this TCE could be a hydraulic bypass of the contaminant plume beneath the PRB.

Monitoring wells placed below the PRB (101A, 103A, 201A, 203A, 301A and 303A in

Fig. 1), however, all show TCE and cDCE concentrations less than detection limits,

confirming that there is likely no hydraulic bypass of the PRB.

Downgradient wells 173, 186 and 700B show significantly different isotopic results

(Fig. 1, Table 1). In contrast to wells 707B and 708, TCE in well 700B has a d13C of

�21.8x, which is enriched in 13C with respect to the average upgradient value

(�24.2x), and to the TCE in the well directly upgradient within Transect 1 (well 101B,

Fig. 2a, �23.8x). This change in the d13C of the TCE, albeit small, is consistent with

typical fractionation associated with degradation whereby a preferential rate of degradation

of the 12C-bonds results in a preferential enrichment of the remaining parent contaminant

(in this case TCE) in 13C (Dayan et al., 1999; Bill et al., 2001; Slater et al., 2002; VanStone

et al., 2004). This is manifested as a less negative d13C value (�21.8x) versus its

presumed starting composition (mean upgradient d13C value of �24.2x). The d13C value

of cDCE in this well also supports this interpretation. When cDCE is produced as a

breakdown product of TCE, it will initially be more depleted in 13C (resulting in a more

negative d13C value) compared to the TCE, reflecting the preferential rate of reaction for12C-containing molecules and preferential incorporation of 12C into cDCE. The d13C value

of cDCE (�29.0x) observed in downgradient well 700B compared to the upgradient

wells indicates that, at least in part, cDCE is a product of TCE degradation. As observed

for TCE, concentrations of cDCE drop to below detection limits within the PRB and in the

well below it (well 101A) (Fig. 1a). Thus, it can be inferred that the contaminant plume is

not bypassing the PRB. As observed for wells 707B and 708, the TCE and cDCE observed

in well 700B are likely due to desorption from local downgradient aquifer materials.

Unlike in wells 707B and 708, however, there may be some microbial degradation

occurring in the localized region around this well to account for the enrichment of 13C

observed in the TCE compared to upgradient wells, and the depletion of 13C observed for

the cDCE relative to the d13C values of the TCE in the same well.


Two other downgradient wells (wells 173B and 186) have significantly enriched d13C

values of TCE (�16.8x and �7.0x, respectively) compared to the upgradient wells,

consistent with much more extensive degradation of TCE at these wells. In addition, the

concentrations of TCE observed in these wells are the lowest in the downgradient plume

(25 to 31 Ag/L), and VC is detected in both wells (1.3 to 62 Ag/L). Both cDCE and VC are

typical products of the reductive dechlorination of TCE. The carbon isotopic compositions

of cDCE and VC provide evidence that they are indeed present due to the degradation of

TCE. The d13C value for cDCE in well 173B (�34.8x) is much more negative than that

of TCE (�16.8x), reflecting the preferential incorporation of 12C into the breakdown

product, cDCE. Similarly the d13C value of cDCE in well 186 is �9.6x, which is more

negative compared to the value of TCE in the same well (�7.0x). In addition, in well

186, VC has a d13C value of �18.5x, a value significantly more negative than both

cDCE and TCE in the same well, and again consistent with progressive incorporation of12C into the breakdown products. Furthermore, the cDCE and TCE in well 186 have

extremely enriched d13C values (�9.6x and �7.0, respectively) compared to the

upgradient mean (�24.2x). No process other than degradation is known to produce such

isotopically enriched values through preferential transfer of 12C to the breakdown products

and a corresponding enrichment of 13C in the remaining contaminant.

Based on d13C values alone, it is not possible to determine if the degradation is

microbially or abiotically mediated. However, microbiological analysis of a water sample

taken from well 186 showed a highly increased anaerobic microbial population in this well

compared to other areas in the vicinity of the PRB, suggesting that in situ anaerobic

microbial degradation may account for the CHC composition in this well (USAF, 2003).

Microbiological data are not available from wells 173B and 700B, however, the degree of13C fractionation in TCE and cDCE observed in these wells, combined with their

concentration trends across these PRB transects provide indirect evidence for local

microbiological dechlorination of previously existing contaminants desorbing from the

aquifer matrix, rather than incomplete degradation in the PRB. For example, the

concentration and carbon isotopic composition of TCE and cDCE for well 173B are

shown for Transect 3 in Fig. 2c. The decrease in concentrations of TCE and cDCE across

the PRB at this location indicates that complete degradation is occurring within the PRB.

Furthermore, the concentrations of TCE and cDCE below the PRB (well 301A) are below

detection limits. The carbon isotopic compositions of TCE (�16.8x) and cDCE

(�34.8x) in well 173B, however, indicate that some degradative process is acting on

the chlorinated ethenes downgradient of the PRB. A likely hypothesis is that localized

microbiological activity is degrading TCE and cDCE desorbing from the aquifer material.

5. Conclusions and implications for PRB technology

Carbon isotopic compositions of TCE in two of the wells downgradient of the PRB

(wells 707B and 708) indicate that the observed downgradient concentrations are likely

due to desorption of contaminant sorbed to the aquifer material prior to the PRB

installation. In contrast, the carbon isotopic results support the existence of microbially-

mediated reductive dechlorination of TCE in well 186. Furthermore, d13C values of TCE


and cDCE in downgradient wells 173B and 700B appear to indicate a local biological

dechlorination of contaminants desorbing from aquifer material, rather than incomplete

degradation in the PRB.

A complex series of processes control concentrations of pollutants at contaminated

field sites. Both abiotic and biotic degradation may reduce contaminant concentrations in

the dissolved phase, but in addition mass transfer processes such as transport by

advection and diffusion and mass partitioning via sorption and desorption play important

roles. Since PRBs are commonly installed within existing contaminant plumes, all of

these processes may play a role in controlling concentrations of contaminants

downgradient of the barrier. Specifically, at installations where downgradient concentra-

tions persist at higher than anticipated levels, the possibility of inadequate treatment

within the barrier, hydraulic bypass around or under the barrier, and desorption of

material from the downgradient aquifer material must typically be assessed. Measure-

ment of the distribution of concentrations alone can be ambiguous. Similarly, as this

study shows, stable isotope measurements alone are also no silver bullet. Nonetheless, it

is clear that integration of both concentration distributions and stable isotope data

provides a significant advantage in terms of constraining the role of processes of

contaminant transformation versus processes of mass transfer in the subsurface. The kind

of process level information stable isotope measurements can provide can be an

important advantage in distinguishing between the effects of fractionating degradative

processes and non-fractionating processes such as desorption. While this study illustrates

these principles at a site contaminated with chlorinated hydrocarbons, the conclusions

are directly applicable to other contamination problems as well, such as abiotic

degradation of carbon tetrachloride (Amonette et al., 2000; Matheson and Tratnyek,

1994; Pecher et al., 2002) which has recently been shown to also involve significant

carbon isotope fractionation (Elsner et al., 2004).

Acknowledgements

The authors gratefully acknowledge the assistance of Stephen Hart and Anna Radloff of

URS Corporation for providing samples and site information. Special thanks are due to the

anonymous reviewers of this manuscript and to Martin Elsner of the Stable Isotope

Laboratory at the University of Toronto for valuable comments. This work was supported

in part by the Natural Sciences and Engineering Research Council of Canada, EnviroMetal

Technologies Inc., and a CRESTech Industrial Cooperative Research Award to N.V.

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