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Journal of Contaminant Hydrology 69 (2004) 233–261

In situ sequenced bioremediation of mixed

contaminants in groundwater

J.F. Devlina,*, D. Katicb, J.F. Barkerc

aGeology Department, University of Kansas, 120 Lindley Hall, Lawrence, KS 66045-7613, USAbGolder Associates, Mississauga, Ontario, Canada

cDepartment of Earth Sciences, University of Waterloo, Waterloo, Ontario, Canada

Received 29 August 2002; accepted 9 July 2003

Abstract

A mixture of chlorinated solvents (about 0.5–10 mg/l), including tetrachloroethene (PCE) and

carbon tetrachloride (CT), together with a petroleum hydrocarbon, toluene (TOL), were introduced

into a 24 m long� 2 m wide� 3 m deep isolated section (henceforth called a gate) of the Borden

aquifer and subjected to sequential in situ treatment. An identical section of aquifer was similarly

contaminated and allowed to self-remediate by natural attenuation, thus serving as a control. The

control presents a rare opportunity to critically assess the performance of the treatment systems, and

represents the first such study for sequenced in situ remediation. The first treatment step was

anaerobic bioremediation. This was accomplished using a modified nutrient injection wall (NIW) to

pulse benzoate and a nutrient solution into the aquifer, maximizing mixing by dispersion and

minimizing fouling near the injection wells. In the anaerobic bioactive zone that developed, PCE, CT

and chloroform (CF), a degradation product of CT, degraded with a half-lives of about 59, 5.9 and

1.7 days, respectively. The second step was aerobic bioremediation, using a biosparge system. TOL

and cis-1,2 dichloroethene (cDCE), from PCE degradation, were found to degrade aerobically with

half-lives of 17 and 15 days, respectively. Compared to natural attenuation, PCE and TOL removal

rates were significantly better in the sequenced treatment gate. However, CT and CF were similarly

and completely attenuated in both gates. It is believed that the presence of TOL helped sustain the

reducing environment needed for the reduction of these two compounds.

D 2003 Elsevier B.V. All rights reserved.

Keywords: Sequenced treatment; Bioremediation; In situ remediation; Natural attenuation; Chlorinated solvents;

Hydrocarbons; Anaerobic–aerobic bioremediation

0169-7722/$ - see front matter D 2003 Elsevier B.V. All rights reserved.

doi:10.1016/S0169-7722(03)00156-6

* Corresponding author. Fax: +1-785-864-5276.

E-mail address: [email protected] (J.F. Devlin).

J.F. Devlin et al. / Journal of Contaminant Hydrology 69 (2004) 233–261234

1. Introduction

The treatment of mixtures of organic contaminants in water has been the subject of study

and practice for many years (Fogel et al., 1982). With the growing acceptance and

implementation of in situ technologies for groundwater remediation, there is an opportunity

to further develop and adapt these systems to the problem of treating mixed contaminant

plumes in the ground. The primary goal of this research was to evaluate the performance of

an in situ groundwater treatment system consisting of two sequenced technologies, the first

anaerobic and the second aerobic. Toward this end, an artificial plume of carbon

tetrachloride (CT, 1–2 mg/l), tetrachloroethene (PCE, 1–2 mg/l) and toluene (TOL, 1–

10 mg/l) was created and treated inside an isolated section of aquifer subjected to a constant

rate of groundwater flow. The treatment system was evaluated against an identical,

adjacently located section of aquifer, in which natural attenuation was permitted to occur.

Numerous investigators have suggested that sequenced in situ treatment of groundwater

should be beneficial (Zitomer and Speece, 1993; Guest et al., 1993; Dickel et al., 1993;

Nelson et al., 1994; Knackmuss, 1996; Agrawal and Tratnyek, 1996; Hofstetter et al.,

1999). Attempts have been made, or are in progress, to implement systems of this kind

(Tratnyek and Johnson, 1998; Lizotte et al., 1999; Barcelona and Xie, 2001). To our

knowledge, there was only one complete and controlled field test of such a system in the

literature at the time of writing (Morkin et al., 2000). In that experiment, an anaerobic–

aerobic treatment sequence was engineered by arranging a granular iron reactive barrier

upgradient from a biosparge unit, within a Funnel-and-Gate structure (Starr and Cherry,

1994). The contaminants of concern included trichloroethene (TCE, 0.09 mg/l), cis-

dichloroethene (cDCE, 198 mg/l), vinyl chloride (VC, 48 mg/l), benzene (0.21 mg/l), TOL

(6.3 mg/l) and traces ( < 0.5 mg/l) of other chlorinated and aromatic volatile organic

substances. The granular iron barrier reduced the concentrations of chlorinated compounds

between 91% and >99%, where cDCE and TCE were the respective end-members of this

range. The biosparge zone removed about 66% of the remaining cDCE and VC from the

groundwater. At the end of 8 months of operation, the system successfully lowered

contaminant concentrations below the target levels, which were based on the California’s

maximum contaminant levels (MCLs). Thus, there is evidence that sequenced in situ

treatment is viable, but additional work is needed to explore alternative technologies in

such sequences.

A variety of in situ technologies for groundwater cleanup have been introduced over the

past two decades. These fall into two classes: passive and active technologies. The passive

technologies, such as permeable reactive barriers (PRB), attempt to maximize the benefit

of low operation and maintenance costs for long-term operation. In these systems, there is

little or no pumping, and treatment is ensured either by the installation of a suitably large

PRB, or through the use of a passive flow-control structure, such as a Funnel-and-Gate

(Starr and Cherry, 1994) or a Trench-and-Gate (Bowles et al., 2000). An example of a PRB

is the granular iron barrier, first proposed by Gillham and O’Hannesin (1992) for the

treatment of chlorinated organics, and later used for the treatment of a variety of other

contaminants (chromium, Pratt et al., 1997; uranium, Morrison et al., 2001; arsenic, Su

and Puls, 2001; nitroaromatics, Devlin et al., 1998; nitrate, Alowitz and Scherer, 2002) and

used in the sequenced system mentioned above (Morkin et al., 2000).

J.F. Devlin et al. / Journal of Contaminant Hydrology 69 (2004) 233–261 235

The active technologies place more emphasis on lower capital costs and, arguably,

greater control over the system hydraulics (achieved by pumping). Enhanced bioremedi-

ation has generally fallen into the category of active in situ technologies, since it often

involves continuous pumping or recirculation of water. Traditionally, the recirculation has

taken place between conventional wells dedicated to either extraction or injection (Norris

et al., 1994). However, recirculation systems using single boreholes have also been tested

and found to be promising in at least one stratified aquifer (McCarty et al., 1998).

Although both passive and active methods have advantages in different situations, it was

decided that this research would focus on the passive technologies. In order to extend our

experience with treatment sequences beyond granular iron PRBs for anaerobic treatment,

anaerobic bioremediation was chosen for inclusion in this experiment.

An approach to bioremediation that minimizes pumping while maximizing dispersive

mixing was tested in the CFB Borden aquifer and reported by Devlin and Barker (1994,

1996, 1999) and Devlin and Muller (1999). Briefly, the method involves the injection of a

nutrient or primary substrate solution from a nutrient injection wall (NIW) at regular

intervals determined by the average linear groundwater velocity. The solution pulses are

timed to that they overlap due to hydrodynamic dispersion at some predetermined distance

(on the order of a few metres) from the injection wells. Where the overlap occurs,

subsurface bacteria receive a continuous supply of nutrients/substrate, and a bioactive zone

is sustained. Although the NIW reported by Devlin and Barker (1996) was successful in

creating and sustaining a modified anaerobic redox zone in which CT was transformed, it

was unsuccessful at stimulating the reduction of TCE (Devlin, 1994; Devlin and Muller,

1999). A challenge for this study was to achieve chlorinated ethene reduction with the

NIW technique.

The biodegradation of aromatic hydrocarbons (e.g. TOL) and lesser chlorinated ethenes

(e.g. cDCE, VC) in aerobic environments is known to occur relatively rapidly (Norris et

al., 1994; Bradley and Chapelle, 1998). Moreover, there are several methods currently

available for introducing oxygen to the subsurface to support aerobic biodegradation.

These include air sparging, sometimes called biosparging (Norris et al., 1994; Morkin et

al., 1998), the use of an oxygen releasing metal peroxide compound (ORC; Regenesis

Bioremediation Products, CA) (Borden et al., 1997; Chapman et al., 1997; Landmeyer et

al., 2001) and oxygen introduction through diffusive emitters (Wilson et al., 2002). At the

time of this experiment, the oxygen diffusive emitters were under development and not

available for use. ORC was tested as part of a sister project (Barker et al., 2000). Therefore

a variation of biosparging was selected as the means of aerating the aquifer downgradient

of the NIW (Barker et al., 1993). Sparging was conducted only to emplace or renew

residual oxygen bubbles in the porous medium, with minimal excess gas delivery. The

residual oxygen bubbles dissolved into the flowing groundwater, providing a readily

available electron acceptor.

An inevitable consequence of sequenced anaerobic and aerobic treatment systems is

that a distinct geochemical boundary is created at the transition between the two systems.

Morkin et al. (2000) did not observe any performance limiting effects at the boundary.

However, their sequence included a limestone gravel buffer zone separating the two

extreme environments. The current design contained no such buffer zone. A further goal of

this work, then, was to evaluate any effects of the anaerobic–aerobic boundary on the


performance of the system, with emphasis on the possible flow interferences that might

result.

2. Site description

The study site has been described previously (Devlin et al., 2002). Briefly, the

experiment was conducted in a pristine, well-sorted, sandy aquifer located at the Canadian

Forces Base Borden (CFB Borden). The aquifer was 3 m deep and underlain by a 6-m-

thick aquitard which was determined to be an effective hydraulic barrier beneath the

aquifer. Borden sand contains a significant carbonate component that tends to buffer the

pH of the groundwater to near neutral values. It contains low organic carbon

(foc = 0.0002), and a porosity of about 0.33 (Ball et al., 1990).

The water table varied between 0.3 and 1.5 m below ground surface (bgs), depending

on time of year and location. A sheet pile wall bounded the test section on three sides,

creating an alleyway measuring 2 m wide� 24 m long, and hereafter referred to as the gate

(Fig. 1). An identical alleyway was constructed immediately next to the current one, and

served as a control; no treatment measures were attempted in that section of the aquifer

(Devlin et al., 2002). The gates were covered with a 30 m long� 9.2 m wide greenhouse

structure consisting of steel tube framing and a flame retardant, vinyl-coated polyester

cover to protect the equipment and instrumentation from the elements. The structure also

facilitated sampling throughout the year and prevented direct local recharge due to rainfall

or snowmelt, which might have influenced the hydraulics.

The groundwater flow inside the treatment gate was maintained by continuous

pumping of a fully screened 5.08 cm (2 in.) well, located at the closed end of the

structure. The pumping rate was maintained at 130 ml/min. The uniformity of the flow

was assessed by a bromide tracer experiment tracked with fences of multilevel monitoring

Fig. 1. Schematic of the treatment gate.


wells (see Section 4). Average linear groundwater velocities were determined to be in the

range of 4–15 cm/day, expressed as fence averages. A value of 14 cm/day was considered

most representative of conditions in the gate and was used in all calculations.

3. Experimental design

As alluded to above, the test section of aquifer was hydraulically separated from the bulk

of the aquifer by sheet piling (Fig. 1). The gate was enclosed on three sides with a pumping

well at the closed end; the fourth side was left open to permit native groundwater to enter

the test section. By isolating the test section in this way, consistency of flow could be

assured so that contaminant transformations could be interpreted with minimal ambiguity.

Three test compounds, CT, PCE and TOL, were introduced to the groundwater stream at the

open end of the gate using diffusive emitters (see below). This design permitted a long-term

source that could be discontinued without excavation at the conclusion of the experiment.

The target concentrations for the contaminants were 1–10 mg/l for TOL, and 1–2 mg/l for

PCE and CT. Previous work established that the retardation factors for these compounds in

the Borden aquifer were about 1.2, 2.7 and 1.2, respectively (Pankow and Cherry, 1996;

Devlin, 1994; Hubbard, 1992). In addition, chloroform (CF) formed from the transforma-

tion of CT; it was previously found to have a retardation factor of about 1.1 in the Borden

aquifer (Devlin, 1994). Also, cDCE was produced from PCE degradation and was

estimated to have a retardation factor of 1.1, based on its reported Koc, and the retardation

behaviour of PCE (Mercer and Cohen, 1990).

A NIW configured to introduce benzoate and inorganic nutrient pulses to the aquifer

was installed immediately downgradient from the source to create an 8-m-long anaerobic

zone for reductive dechlorination. The amendment pulses were added at intervals of 20–

30 days to support a bioactive zone about 4 m downgradient of the NIW wells. At the

outset of the experiment, benzoate was the sole ingredient in the amendment solution. It

was prepared as a concentrated solution of sodium benzoate, 140 g benzoate/kg solution,

and added directly to the recirculating water from the NIW, in-line. Once mixed with the

groundwater, the benzoate in the NIW was diluted to about 250 mg/l. On day 151 (after

contaminant additions began), the benzoate concentrations were reduced to 125–160 mg/

l without adversely affecting the anaerobic zone. On day 207 and for each injection

afterwards, the amendment solution was supplemented with nutrients consisting of a

mixture of approximately 10 mg/l of K2HPO4, KH2PO4 and NH4NO3, as well as 2 mg/l of

MgSO4�7H2O.

The interval between amendment pulses was determined using velocity and dispersion

parameters estimated from a bromide tracer pulse included in one of the benzoate pulses.

Using these transport parameters, a 1D solution to the advection–dispersion equation

(Devlin and Barker, 1996) was used to estimate the dimensions of a pulse that had been

transported 5 m from the NIW. The 5-m distance corresponded to the target location for

the beginning of the anaerobic bioactive zone. A pulsing period of 28 days was determined

appropriate from this analysis.

A residual oxygen barrier (ROB) was installed downgradient of the NIW. By placing

the aerobic treatment zone after the anaerobic treatment, partially dechlorinated inter-


mediates of PCE, such as cDCE and VC, could be further degraded at the same time as

TOL. Moreover, the aesthetics of the water emerging from the gate would be better after

aeration.

Sampling was conducted using multilevel wells to provide detailed vertical and lateral

data for the assessment of contaminant fate. Details of the installation and configuration

are provided below. The experiment was active for a period of about 1 year.

4. Methods

4.1. Site instrumentation

Waterloo BarrierR sealable joint sheet piling was driven into the ground and keyed into

the aquitard to break any hydraulic connection between the inside of the gate and the

surrounding aquifer. The joints were filled with grout, following installation, as a further

precaution against contaminant leakage out of the test volume. The integrity of the seals

was assessed throughout the experiment, and found to be satisfactory, by regular sampling

of monitoring wells located immediately outside the gate.

The organic test compounds, CT, PCE and TOL, were released from diffusive emitter

devices (Barker et al., 2000; Arildskov and Devlin, 2000). These devices have the

advantage of creating continuous sources for whatever period of time is required. When

the source is no longer needed, it can be stopped with little time and trouble. The emitters

were installed in four fully screened 20-cm-diameter PVC wells located at the open end of

the gate. Briefly, they were designed to continuously release about 1 mg/l of each

contaminant (CT, PCE, TOL) into the aquifer.

The emitters performed adequately for the test, although during the time period

between days 75 and 150, PCE, CT and TOL input concentrations (as fence 302 averages)

rose unexpectedly. In general, PCE was introduced at approximately 200–300 Ag/l, exceptbriefly near day 150 when inputs rose to 600 Ag/l. CT was introduced at about 200–600

Ag/l, except between days 100 and 150 when the concentration rose to 2000 Ag/l. TOL was

introduced at about 400–1000 Ag/l, except between days 75 and 110 when concentrations

rose to about 10,000 Ag/l. Despite these difficulties, concentrations further along the flow

path (fence 306), upgradient of any transformations, remained quite steady (PCEf 100–

140 Ag/l; TOLf 70–140 Ag/l; CT not assessed at 306 due to transformation to CF),

indicating that the source anomalies were of minor importance to the plume. Further

details of the source emitter performance can be found in Barker et al. (2000) and Devlin

et al. (2002). Improvements to the emitter design were reported by Arildskov and Devlin

(2000).

Multilevel monitors were installed across the gate, in fences, at distances of 0.8, 9.29,

11.37, 13.08, 17.29, 18.66, 19.3, 20.4 and 22.65 m from the source wells, and were

designated G302, G310, G312, G315, G318, G320, G321, G322 and G324, respectively

(Fig. 1). Each fence consisted of three monitors designated L, M or R depending on

whether the device was located on the left middle or right side of the gate (looking

downgradient). Each multilevel was constructed from four 0.3-cm outer diameter (OD)

stainless steel (ss) tubes strapped to a central 1.25-cm OD PVC stalk. The ss tubes


terminated at depths of 1.2, 1.9, 2.6 and 3.3 m bgs, designated levels 1, 2, 3 and 4,

respectively. No filter packs were used in the completion of the multilevel monitors. Each

monitoring point was completed with a fine ss mesh wrapped around the ends to prevent

invasion by the native sand.

About 5 m into the gate, the NIW was installed using the materials and methods

described in Barker et al. (2000) and Devlin and Barker (1999); the NIW section was

excavated, instrumented and backfilled with a coarse filter sand. Three injection and

withdrawal wells were installed in two rows (Fig. 2). All were fully screened PVC

construction, and 5.08 cm in diameter. Nutrient amendments were accomplished using a

closed loop dosing system as described in Devlin and Barker (1996). However, in contrast

to the previous NIW, nutrient pulses were emplaced by flushing the barrier in the direction

of ambient groundwater flow, rather than perpendicular to it. In a gate setting (as opposed

to an open aquifer), this configuration was preferred since it reduced the effects of

asymmetric nutrient distribution, due to flow convergence at the pumping wells.

The ROB, located about 16 m into the gate, consisted of six horizontal sparge wells

installed at the base of the aquifer. Each device was perforated in six evenly spaced

Fig. 2. Schematic of the NIW.


locations along the horizontal tubes. The wells were oriented perpendicular to the flow

direction and were equally spaced over a 2.8-m-long section of the gate (Fig. 3). Each

sparge well contained four 0.62-cm-diameter polyethylene tubes terminating at different

distances along the horizontal pipe. The tubes were pressurized simultaneously to ensure

that oxygen delivery occurred as uniformly as possible. Additional details of the

construction of the sparge system are given in Barker et al. (2000). The sparge wells

were overlain by a gravel fill with a porosity of 0.46. The upper 30 cm of the ROB

section was backfilled with pea gravel (grain diameters 0.5–1.4 cm) along the central

axis of the gate and filter sand along the sides. The unsaturated zone, to a depth of 22

cm, was monitored with four 0.3-cm inside diameter (ID) ss soil gas sampling tubes

equipped with mininert caps. These were installed in the pea gravel along the central

axis of the gate. The ROB was covered at surface with a plastic tarp to minimize the

escape of volatile compounds during sparging. Escape laterally was controlled by

installing partially penetrating sheet piles (extending to 1 m bgs) on either end of the

ROB.

The horizontal sparge wells were constructed from 5.1-cm outer diameter (OD) PVC

pipes installed vertically, three on each side of the gate, from the surface to the bottom of

the ROB where they were connected to same sized horizontal pipes extending across the

gate with six vent holes each. Inside each pipe were four 0.62-cm OD polyethylene tubes

running from surface to corresponding vent holes. At surface, the tubes were connected to

a manifold that was itself connected to a cylinder of oxygen. Gas injections were

performed at regular intervals using a household digital timer from Noma (model

Fig. 3. Schematic of the ROB.


N1507) to control a two-way, direct lift diaphragm solenoid valve (from Cole–Palmer)

that was installed between the oxygen cylinder and the manifold.

Test sparging was completed once the ROB was backfilled with gravel up to the level

of the water table. These tests indicated that a 1-min sparge at 4–5 psi of gauge pressure

was sufficient to establish a bubble residual over about 60% of the barrier and to begin gas

transfer to the unsaturated zone. Based on this, and a gas filled porosity of 4% of the

sparged aquifer volume (determined in laboratory experiments (Barker et al., 2000)), it

was estimated that each 1-min sparge released a maximum of about 400 l of oxygen (3.0 m

deep� 2.0 m wide� 2.8 m long� 0.04 porosity� 0.6 coverage; note: 3 m deep corre-

sponds to the water table at surface—representing the maximum saturated aquifer

thickness), corresponding to 22 mol. With a benzoate loading of about 0.46 mol/day (a

maximum of 300 mg/l entering the gate on average at 130 ml/min) and the following

stoichiometry for aerobic benzoate consumption

C7H5O�2 ðaqÞ þ

15

2O2 ðaqÞ þ Hþ

ðaqÞ ! 7CO2 ðaqÞ þ 3H2O ðlÞ

it was calculated that (22 mol O2 per sparge/0.46 mol benzoate per day/(15/2)=) 6.5 days

between sparges would be the interval necessary to meet the oxygen demand. A

conservative semiweekly sparging schedule (3.5-day interval) was adopted at the outset

of the experiment; as discussed below, this was later modified to daily, minute-long

sparges.

Immediately next to the gate was a second three-sided sheet pile alleyway of identical

dimensions and similarly instrumented with monitoring wells. However, this structure

contained no NIW or ROB, but served instead as a control so that degradation resulting

from bioremediation could be compared to degradation due to natural attenuation

processes. Details of the control gate construction and performance are described

elsewhere (Devlin et al., 2002).

4.2. Sampling methods

Sampling was conducted for a variety of solutes and geochemical parameters.

Complete details of the protocols and QAQC are reported in Barker et al. (2000). Briefly,

all aqueous samples were collected from the multilevel points using either a peristaltic

pump or a 60-ml syringe for suction; the shallow water table and permeable aquifer

material contributed to minimal degassing of samples. The monitors were purged of a

volume equivalent to three times the standing water volume of the deepest point (57 ml)

prior to collecting samples. Dissolved organic solute samples, including volatile organics

compounds (VOCs) and organic acids, were collected in 40-ml screw cap glass vials with

Teflon-lined septa. The vials were placed in line before the pump to avoid contact with

pump-head tubing.

Inorganic samples were stored in polypropylene bottles; cations were filtered with 0.45-

Am membranes and acidified with nitric acid (10 drops per 60 ml bottle) to achieve a pH of

about 2. Anions were collected unfiltered and unacidified. Dissolved oxygen concen-

trations were measured either with an Orion model 835 meter and a model 083010 probe


(for concentrations >0.5 mg/l), or using the Chemetrics colorimetric method (for concen-

trations < 0.5 mg/l).

Hydrogen was measured using the method of Chapelle et al. (1997). Three samples

were collected from any given location and the results averaged for later interpretation.

4.3. Analytical methods

Nonchlorinated hydrocarbon gases were analysed using an HP 5840A gas chromato-

graph (GC) equipped with a 2-ml sample loop, a 30-m Megabore GS-Q capillary column

with helium flowing at 12 ml/min serving as the carrier gas and an oven temperature of

100 jC. The injector temperature was 100 jC and the detector, a flame ionization detector,

was kept at 200 jC. VOCs were analysed using an HP 5890 GC equipped with a

photoionization detector and a Varian Genesis headspace autosampler.

Inorganic anions were analyzed commercially by Water Technology International

(WTI), Burlington, Ontario, using a Waters Ion Chromatograph equipped with a Waters

Series 590 pump, a WISP UK Rheodyne Sample Injector, an IC-Pak A HR column, a 430

conductivity detector and a Spectra Physics SP4290 integrator. Cations were also analyzed

at WTI, by inductively coupled plasma (ICP).

Method detection limits for VOCs were estimated from calibration curves as

described by Devlin (1996). They were under 10 Ag/l for all compounds identified in

this study. Precision was estimated by comparing the analyses of duplicate pairs. In most

cases, precision was in the range of F 15–25% (including methane). The exceptions

were PCE (F 30%) and carbon tetrachloride (F 75%). In the latter case, the high

percentage uncertainty is reflective of the low concentrations in the duplicates collected

during the QAQC program, rather than the overall quality of the analyses. Even so,

since carbon tetrachloride transformed rapidly and was not present in most of the gate,

the uncertainty experienced in this compound’s analysis was not an obstacle to

interpretation.

Inorganic anions and major cations all had method detection limits of 1 mg/l or less,

and were analyzed with precision of F 10% or less, except for chloride (F 15%) and

dissolved oxygen (DO, F 40% at concentrations less than 0.5 mg/l). Since the

concentrations of chlorinated compounds introduced to the gate were quite low in

comparison to background chloride, there was little chance for a chloride mass balance

even with a very sensitive and precise analytical technique. Therefore, chloride was

examined for site characterisation purposes only, and the precision experienced was

considered acceptable. DO levels below 0.5 mg/l were considered indicative of bulk

anaerobic conditions; no attempt was made to use the measured DO levels for more

quantitative purposes.

4.4. Data reduction

Mass balance calculations were performed using the FORTRAN program FENCE

(Barker et al., 2000) to integrate breakthrough curves for point specific breakthrough

masses, and then adding the masses to obtain totals for each fence. Flux estimates were

obtained by dividing these mass totals by the cross-sectional area (6 m2) and the number of


days over which the masses accumulated. Average linear velocities were estimated from

the arrival times of the centres of mass of the breakthrough curves.

5. Results and discussion

5.1. Tracer test

The tracer mass at the various monitoring fences (Fig. 1) was found to balance to

withinF 30%, in agreement with a similar tracer test conducted in the control gate (Devlin

et al., 2002). Velocities were estimated to be in range of 4–14 cm/day depending on the

sampling point examined. A subsequent tracer test, in which bromide was released from

the NIW, indicated that the average velocity was closer to 15 cm/day during the period

data were collected from the anaerobic treatment zone. A value of 14 cm/day was used in

the calculations presented here.

The bromide tracer test was used to characterise longitudinal dispersion and velocity

(14 cm/day) downgradient of the NIW. This was accomplished by averaging instantaneous

concentrations at fence G312, and fitting the resulting averaged breakthrough curve with a

1D solution of the advection–dispersion equation (Devlin and Barker, 1996). The fence-

averaged dispersivity was found to be 0.21 m, corresponding to relatively low dispersion

in the gate. In fact, degradation rate constants calculated from equations that included

dispersion and those from equations that neglected dispersion were less than 10% apart

(Devlin et al., 2002). This justified a simplified analysis of the data and the omission of

dispersion terms where convenient. Although, in this case, dispersion could be neglected

in the estimation of degradation rate constants, it was still occurring and still important for

the proper operation of the NIW.

5.2. Redox control using the NIW and ROB

Changes in redox conditions in various parts of the gate were tracked by monitoring

several redox sensitive parameters including DO, Fe(II), SO42�, H2 (aq) and Eh. Prior to

beginning the experiment, the aquifer was characterised as being aerobic, with background

DO of 2–7 mg/l (Devlin et al., 2002). Nitrate was not detected in the groundwater and iron

concentrations were below 0.5 mg/l. Sulphate was present at concentrations of 15–20 mg/

l and Eh measurements were in the range of 50–300 mV. However, traces of methane

were detected in samples (maximum of 36 Ag/l) suggesting the presence of microenviron-

ments that were more reducing than the bulk of the aquifer. In addition, hydrogen

concentrations measured in background wells yielded values of 0.2 nM, suggesting that

iron reducing conditions prevailed.

By day 220, there were signs in the deepest parts of the aquifer, corresponding to

multilevel depths 3 and 4 (2.5–3.3 m bgs, respectively), that conditions had become more

reducing as a result of the benzoate additions. Sulphate was no longer detectable at these

lower depths and iron concentrations had increased from < 0.5 mg/l to an average of about

1 mg/l at G312. Moreover, methane concentrations increased to between 1000 and 2000

Ag/l in selected locations (e.g., G312M2, G312M4, G315M4) between G312 and G315.


Additional sampling on day 339 indicated that the trend toward increasingly reducing

conditions between G310 and G315 was continuing. Fence-averaged iron concentrations

rose to between 1 and 2.5 mg/l while corresponding sulphate concentrations declined to

< 1 mg/l (averaging all depths). Methane production increased noticeably with concen-

trations between 6000 and 8000 Ag/l in the deepest parts of the aquifer (3.3 m bgs) and

1000–2000 Ag/l at level 2 (1.9 m bgs). Between days 332 and 334, hydrogen measure-

ments were made yielding concentrations between 0.4 and 5.3 nM, the highest measured

during the experiment. The mean hydrogen concentration between G310 and G315 was

3.0 nM, suggestive of sulphate reducing conditions, but the increasing methane concen-

trations indicated that methanogenesis dominated some zones within the aquifer, partic-

ularly at depth.

The groundwater in the ROB was initially aerobic, as described above. Between days 5

and 240 the ROB was sparged semiweekly. During this time, DO levels in the gravel fill

varied from below detection to 6 mg/l. On day 238, gas flow rate measurements were

made between the oxygen tank and the wells. It was discovered that each sparge was only

introducing 35 l of oxygen to the gravel fill, far less than the maximum 400 l anticipated in

the design. Increasing sparge pressure was ruled out as a remedy because it was feared that

this might cause large channels to form in the gravel, reducing the volume of porous

medium receiving oxygen bubbles (Ji et al., 1993; Austrins, 1997). Instead, it was decided

to increase the sparging frequency to once daily, for 1 min. Once this procedure was

adopted, on day 240, dissolved oxygen concentrations in the groundwater increased to

between 10 and 40 mg/l at most monitored points in the ROB. Exceptions included the

deepest monitoring points (levels 4) where DO was between 2 and 10 mg/l (note that these

points were located below the elevation of the sparge nozzles in the horizontal wells) and

at monitor G316R, all depths, where DO levels were found in the range 1–12 mg/l. On the

basis of these data, the ROB section was determine to be aerobic for the last 100 days of

the experiment.

Downgradient of the ROB, oxygen was rapidly depleted due to the presence of

benzoate from the NIW, TOL in the plume, and acetate produced in the anaerobic

treatment zone through fermentation of the two preceding compounds. Hydrogen levels in

the aquifer downgradient of the ROB were consistently 0.1–0.7 nM, suggesting an iron

reducing environment. This finding was consistent with an increase in Fe(II) concen-

trations (from < 0.5 to about 1 mg/l) in the groundwater sampled from G318 to G320

between days 220 and 238.

5.3. NIW treatment of VOCs

The test contaminants, CT, PCE and TOL, were released from the source wells and

permitted to migrate through the NIW into the anaerobic treatment zone. CT was rapidly

transformed to CF before the NIW, so for all practical purposes, CF became a source

compound rather than CT. This rapid transformation was also observed in the control gate

(Devlin et al., 2002).

Benzoate and, beginning on day 207, inorganic nutrient additions were pulsed into the

aquifer an average of once every 25 days (Katic, 1999) and the procedure required about

5 h to complete (0.21 days). So, the NIW system operated passively 99% of the time.


Residence times in the NIW varied from 22 days (R = 1.2, estimated for TOL and cDCE)

to 50 days (R = 2.7, estimated for PCE), assuming a groundwater velocity of 14 cm/day.

Thus, the more mobile compounds, TOL and CF, were likely recirculated at least once

during pulsing before passing into the anaerobic treatment zone while PCE may have

been recirculated twice, assuming similar mobility in the filter sand and the aquifer sand.

The recirculation apparatus was fashioned in a closed loop, so losses due to volatilisation

were considered unlikely; DO concentrations remained unchanged in the NIW before and

after injections, further indicating limited contact between the groundwater and the

atmosphere. Sorptive losses to pipes and flowmeters are likely to have been minor after

the first injection because the flow rates were high (several gallons per minute) and

sorptive uptake by the plastics used in construction would be expected to slow with

increased contact time (Parker and Ranney, 2000; Arildskov and Devlin, 2000). The

effect of recirculation on maximum contaminant concentrations is thought to be primarily

one of dilution due to mixing with water from less contaminated zones, with the result

being an increased uniformity of contaminant concentrations throughout the NIW. For

this reason, the analyses that follow will be limited to data from and downgradient of the

NIW.

In order to anticipate the progress of the contaminants as they were transported

through the gate, a 1D solution to the advection–dispersion equation was used that

accounted for linear equilibrium retardation; reactions were not considered in these

simulations (Fig. 4). Transport parameters were taken from the results of the preceding

tracer tests. This preliminary modeling indicated that by day 238, the mobile compounds,

TOL and CF, should have arrived at the pumping well, while PCE would have arrived at

the ROB. By day 339, TOL and CF were predicted to be present throughout the gate at

input levels, while PCE was anticipated to be near input levels in the ROB, and just

beginning to break through at the pumping well. Comparing these predictions to the

fence-averaged data from the experiment, it can be seen that TOL levels (f 0.9 mg/l)

were indeed quite consistent between the NIW and ROB on day 238 (samples were not

yet routinely collected beyond the ROB at this time) and PCE appeared to arrive at the

ROB at this time, in accordance with predictions (Fig. 5). No PCE degradation products

(e.g., cDCE, TCE) were detected in the groundwater between days 1 and 238, suggesting

that up to this time PCE degradation was minimal. On the other hand, CF concentrations

in the experimental gate were considerably less extensive than predicted. On day 112,

CF had been detected at the monitoring fence 11.5 m from the source wells in

concentrations in excess of 100 Ag/l, but by day 238, CF was absent from the anaerobic

zone. Since the source wells remained active during this time, these data indicate that CF

underwent degradation in the anaerobic treatment zone. No chlorinated methanes other

than CF or CT were identified in the groundwater, suggesting that complete dechlori-

nation was occurring.

By day 339, PCE had not measurably advanced, and may even have declined in

concentration within the anaerobic treatment zone (Fig. 5). At the same time, cDCE

appeared upgradient and within the ROB, and TCE was detected in trace quantities ( < 5

Ag/l) upgradient of the ROB (data not shown). TOL levels declined slightly in the

anaerobic zone between days 238 and 339, and both CT and CF remained below detection

at all but one sampling point (G310R, 184 Ag/l CF). These data suggest that while TOL

Fig. 4. Calculated relative positions of PCE and TOL on days 238 and 339, assuming an average groundwater

velocity of 15 cm/day, and a dispersivity of 0.2 m. Retardation factors were 1.2 for TOL and cDCE and 2.7 for

PCE.


degradation was minimal in response to the NIW amendments, both CF and PCE were

being reductively dechlorinated at the end of the experiment.

The first day cDCE was detected in the treatment gate was day 316. At that time, it was

present in highest concentrations (40–70 Ag/l at several monitoring points) at G315,

immediately upgradient of the ROB. The compound was also detected at monitoring

fences downgradient of the ROB, including G318 (f 30 Ag/l at several points). SincecDCE is a reduction product of PCE, it must have been formed in the anaerobic section of

the gate and migrated into the aerobic section. G315 and G318 are about 6 m apart,

Fig. 5. Fence-averaged concentrations of TOL, PCE and cDCE in the treatment gate on days 238 and 339.


corresponding to a travel time of about 48 days (assuming RcDCE = 1.2, as in the NIW

section). So, if cDCE first appeared at G318 on day 316, then it must have been present at

G315 (316� 48 = 268) on about day 270. Given that PCE first arrived at G315 around day

196, without cDCE, the lag time for reductive dechlorination was on the order of

270� 196 = 74 days. This must be considered a minimum since PCE may have been

present at G315 before day 196, at some depths and locations in the aquifer.

5.4. Anaerobic degradation rates

In order to quantify the disappearance rates of CT, CF, TOL and PCE, concentration

profiles and flux calculations were performed. Because CT and CF transformed so quickly


in the gate, there were insufficient data available from which to calculate time integrated

fluxes. Therefore, degradation rate estimates of these two compounds were based on

concentration profiles. CT and CF concentrations at fences G302, G310 and G312 were fit

to the following equations which neglect dispersion and assume pseudo-first-order

transformation kinetics (Devlin et al., 2002) (Fig. 6).

C ¼ Coexp � k

vx

� �ð1Þ

D ¼ FkCo

k � kexp � k

vx

� �� exp � k

vx

� �� ð2Þ

where C = concentration of CT (M/L3), Co = input concentration of CT (M/L3), D = con-

centration of CF (M/L3), k = pseudo-first-order rate constant for CT degradation (T� 1),

k = pseudo-first-order rate constant for CF degradation (T� 1), F = conversion factor that

accounts for reaction stoichiometry (assumed 1 for these calculations) and differences in

molecular weights between parent and daughter compounds when data are reported in mg

rather than mmol ( =WCF/WCT = 119.377 g/mol/153.822 g/mol = 0.77) (dimensionless) and

x = distance from source wells (L). The transformations were largely complete by fence

G310, so the profile data used to fit the equations were sparse. The actual profiles may be

less advanced than represented here and therefore the actual rate constants may be higher

than the estimates. Eq. (1) was first fit to the CT data to obtain Co and k and then Eq. (2)

Fig. 6. Profiles of CT and CF on day 339. Solid circles are observed, fence-averaged CT concentrations, and open

diamonds are observed, fence-averaged CF concentrations. Lines represent modelled fits using Eqs. (1) and (2)

(see text).


was fit to CF data to obtain k (Table 1). The estimated rate constants correspond to half-

lives of 5.9 and 1.7 days for CT and CF, respectively. The CT degradation rate was

comparable with that observed by Devlin and Muller (1999) (t1/2 = 3–7 days) in a separate

experiment at the same site. However, Devlin and Muller did not observe CF degradation

in their experiment. This difference in CF persistence was likely due to the differences in

experimental design. Devlin and Muller introduced CT (which rapidly transformed to CF

and carbon disulphide) into the aquifer as a single pulse. The CF was not in contact with a

given section of aquifer long enough for microbial adaptation to occur and so it was not

degraded (note: there were a small number of locations, corresponding to finer grained

lenses, where methylene chloride was detected in that experiment. Thus, some CF

transformation did occur, but it was minimal). In the current test, the contaminants were

injected continuously. We speculate that this led to more adaptation time and a more

uniformly distributed and complete degradation of CF.

To assess transformation rates for PCE and TOL, fluxes of these compounds in the

anaerobic treatment zone were calculated for fences G306, G310, G312 and G315 for days

154–339 (G306) and 196–339 (all other fences). Previous work in the control gate

demonstrated the utility of this method, and indicated that flux-based rate constant

estimates were likely to be as representative, or more so than those from profile analysis

(Devlin et al., 2002). As described in Barker et al. (2000) and Devlin et al. (2001, 2002),

the equations to estimate a pseudo-first-order rate constant from flux data, neglecting

dispersion, are

lnðJÞ ¼ lnðJoÞ �k

vx ð3Þ

J daughter ¼ J parento F1

k � kkexp � k

vx

� �� kexp � k

vx

� �� exp � k

vx

� �� ð4Þ

where J = flux of solute (M/L2/T) at a point with distance x from the source, and Jo = input

flux of solute (M/L2/T). Eq. (3) was applied to TOL and PCE data (Fig. 7), while Eq. (4)

Table 1

Summary of properties rate constants and half-lives for contaminant degradation in the NIW, ROB and by natural

attenuation

Compound H NIW k NIW t1/2 x ROB

profile k

ROB

profile t1/2

ROB

flux k

ROB

flux t1/2

Nat.

Att. k

Nat.

Att. t1/2

CT 0.12 5.9 0.063 11

CF 0.41 1.7 0.069 10

PCE 0.012 59 –

TOL 0.26 – 1.75e� 03 3.38e� 02 20.5 4.08e� 02 17.0 0.011 63

cDCE 0.074 – 5.05e� 04 6.97e� 02 9.9 4.51e� 02 15.4

METH 22.27 5.29e� 02 0

H reported for approximately 10 jC, cDCE value from Gossett (1987), TOL value from Mackay et al. (1999), and

METH value from Wilhem et al. (1977).

Sparging rate was 3.58� 10� 2 m3 day� 1, the gas filled porosity was estimated to be 3.6� 10� 2, and the volume

of the ROB was 12.3 m3 in the calculations. – = k not distinguishable from 0.

Fig. 7. TOL and PCE flux as a function of distance in the anaerobic section of the treatment gate. Note

logarithmic scale. Field flux values were determined in mg/m2/day. Solid lines are best fits and broken lines

indicate standard error intervals.


was modified, as described below, and used with cDCE data. No significant degradation of

TOL was observed in the anaerobic zone, while PCE was found to be degrading with a

half-life of between 42 and 104 days (range represents F 1 standard error on the rate

constant) with a best fit value of 59 days (Table 1).

The expected degradation sequence for PCE is PCE!TCE! cDCE!VC! non-

chlorinated hydrocarbons (VC refers to vinyl chloride). However, the application of Eq.

(4) to cDCE assumes a degradation sequence of PCE! cDCE! other unidentified

products. Since TCE was only present in trace quantities ( < 5% of the total mass), it

seemed reasonable to treat it as a short-lived intermediate, so the simplified sequence is

justified. Also, since VC was not detected at any time in the gate, it may be

conservatively assumed that cDCE did not further transform, i.e., k = 0. With this

simplification, Eq. (4) becomes

J daughter ¼ J parento F 1� exp � k

vx

� �� ð5Þ

where F is the stoichiometric factor 1 multiplied by the value WcDCE/WPCE=(96.943 g/

mol)/(165.833 g/mol) = 0.585, if PCE degrades along the assumed sequence. It should

be noted that PCE is known to degrade along at least one additional sequence that

involves the formation of chloroacetylene and acetylene as intermediates (Roberts et

al., 1996; Semadeni et al., 1998). Therefore, F could be used as a fitting parameter to

assess the likelihood that PCE was degrading according to a branched pathway as

opposed to a linear one. This possibility is explored below.

The cDCE flux calculated from measured data came to 1.3 mg/m2/day (Fig. 8). From

Fig. 7, JoPCE was estimated to be 3.72 mg/m2/day at G306 (marking the beginning of the

Fig. 8. Predicted and calculated fluxes for PCE and cDCE. Open circles are observed values of PCE fluxes; solid

diamond is the observed cDCE flux at G315. Lines represent model fits.


anaerobic treatment zone) and k/v = 7.78� 10� 2 m� 1. At G315, the only fence where

fluxes of cDCE could be calculated, x= 8.15 m from G306. If it is assumed that F = 0.585

(linear pathway), then from Eq. (5), JcDCE = 1.0 mg/m2/day. The agreement between the

calculated value and the independently determined value, plotted in Fig. 8, is considered

quite good, and supports the hypothesis that PCE degraded primarily to cDCE up to day

339 when the experiment ended.

5.5. ROB treatment of VOCs

Although the DO concentrations were not maintained at the levels desired between

days 1 and 240, as described in Section 5.2, an examination of the TOL profile for

day 238 (Fig. 5) suggested that some aerobic degradation was taking place. This

process was apparently enhanced by day 339, since TOL concentrations in and

downgradient of the ROB declined between days 238 and 339. TOL was in

competition for DO with benzoate, left over from the NIW amendments, and acetate

which formed from benzoate fermentation (Barker et al., 2000). It is believed that a

reduction in benzoate inputs at the NIW could have improved TOL degradation

without seriously affecting the anaerobic treatment. Further work is needed to assess

this possibility.

Aerobic activity was also indicated by the limited penetration of cDCE into the

aerobic treatment zone (Fig. 9). However, since the ROB was sparged regularly, there

remains some question about the role of volatilisation in the declining concentrations of

volatile compounds. In order to assess the rate of mass loss due to volatilisation during

sparging, gas samples were collected from the soil gas monitors installed in the

unsaturated gravel of the ROB. No contaminants were ever detected in these samples.

To further assess the importance of volatilisation, theoretical calculations were performed

Fig. 9. Fence-averaged concentrations of cDCE at G315, immediately upgradient of the ROB, and G318,

immediately downgradient of the ROB. The dissimilar trends suggest attenuation across the ROB.


beginning with the following mass balance which applies to a continuously sparged,

well-mixed batch reactor:

Mw;t þMg;t ¼ Mw;t�Dt þMg;t�Dt �Mg out ð6Þ

where Mw,t =mass of compound in the groundwater at time t, Mw,t� Dt =mass in the

water at time t�Dt, Mg,t=mass in the residual gas phase at time t, Mg,t� Dt=mass in the

residual gas phase at t�Dt, Mg out =mass volatilized due to sparging. Expanding each

term in Eq. (6), invoking Henry’s law (Cg =HCw), and expressing the result in

differential form,

dCw

dt¼ � QgH

hwVtot � HhgVtot

Cw ð7Þ

where Cw = concentration at time t (M/L3), H =Henry’s law constant (dimension-

less =Cg/Cw), Cg = concentration in the gas (M/L3), Vw = hwVtot = volume of water

(L3), Vg = hgVtot = volume of gas (L3), hw =water filled porosity (dimensionless), hg = gasgas filled porosity (dimensionless), Vtot = aquifer volume (L3) and Qg = gas flow rate (L3/

T). Eq. (7) shows that the loss of volatile compounds due to sparging can be

approximated as a first-order kinetic process with rate constant, x, equal to

x ¼ QgH

hwVtot � HhgVtot

:


This first-order decay process can be added to the steady-state advection–dispersion–

reaction equation to account for water flow through the ROB.

BC

Bt¼ 0 ¼ D

B2C

Bx2� v

BC

Bx� kC � xC ð8Þ

If dispersion is once again neglected, the solution to Eq. (8) is a modified version of

Eq. (1).

C ¼ Coexp � k þ xv

x

� �ð9Þ

Theoretical values for x are presented in Table 1, and model profiles were fit to the

fence-averaged data for day 339 of the experiment (Fig. 10). Co was assumed to be the

fence-averaged concentration at G315, and x was measured from G315. The sole fitting

parameter for this exercise was k, the pseudo-first-order biodegradation rate constant.

Profiles were fit to three compounds: TOL, cDCE and methane (METH). The latter

substance was a by-product of the NIW treatment and was included in this analysis

because of its high volatility. If losses due to sparging were significant for any of the

compounds studied, it should be most significant for METH. The magnitudes of k for

the various compounds are discussed in the next section. For our purposes, it is

sufficient to note that when k = 0 in Eq. (9), neither TOL nor cDCE showed significant

concentration declines of importance through the ROB (Fig. 10).

On the other hand, METH was predicted to decline noticeably due to volatilisation, and

the METH profile can be entirely explained by volatilisation alone. However, the

modeling assumed perfect stripping efficiency and continuous gas flow; it seems unlikely

that volatilisation could have occurred in so ideal a fashion. Therefore, the mass losses (of

all three compounds) due to volatilisation are likely to be over-predicted. This has little

effect on the interpretation of the TOL or cDCE data, but some component of the METH

loss may in fact be due to consumption by bacteria.

5.6. Aerobic degradation rates

The profiles were fit with Eq. (9) and pseudo-first-order biodegradation rate constants

(k) were calculated (Table 1). TOL and cDCE were found to be degrading with half-lives

of about 20 and 10 days, respectively. Due to the high Henry’s law constant of METH and

moderate scatter in the profile data (Fig. 10), biodegradation could not be quantitatively

distinguished from volatilisation for this compound.

In addition to the profile data, biodegradation rates were determined from contaminant

fluxes through the gate (Table 1). Fluxes of TOL and cDCE were calculated for the period

after day 238 (once the sparging was occurring daily) and fit with Eq. (3) (Fig. 11). The

TOL degradation rates compare quite well between the two methods used, with half-lives

in the range 17–20 days, suggesting that near steady-state conditions did exist for this

compound. The agreement for cDCE degradation rates is somewhat less remarkable, with

a flux derived half-life of 15 days compared with a profile half-life of 10 days. Since the

flux values are time-integrated, from day 238 to 339, while the profile data correspond to

Fig. 10. Profiles of TOL, cDCE and METH across and downgradient of the ROB on day 339. Broken line

indicates predicted profile if volatilisation were the dominant attenuation process. Solid lines represent best fits

using Eq. (9). Symbols are fence-averaged concentrations.


Fig. 11. TOL and cDCE flux as a function of distance from G315 (near the beginning of the ROB). Symbols

represent observed data; solid lines are best fits using Eq. (3) and broken lines indicate the standard error range.

Fluxes were calculated with the units mg/m2/day.


day 339 only, the disparity might be an indication that cDCE was being degraded more

quickly at the end of the experiment than at earlier times.

5.7. Performance assessment and comparison to natural attenuation

This experiment was conducted immediately next to an identical gate in which no

treatment measures were undertaken. This provided a rare opportunity to compare an in

situ treatment scheme against natural attenuation under virtually identical conditions. A

glance at the distributions of contaminants on day 339 of the experiment reveals that in

both cases CT and CF were attenuated in relatively short travel distances. While CT

degradation occurred almost immediately in both gates, the onset of CF degradation was a

few months earlier in the treatment gate. Differences between experimental and control

half-life estimates are probably not significant given the limited number of data points

available for the treatment gate estimates (Fig. 12).

Perhaps the biggest differences between the two gates was apparent for PCE and TOL,

both of which migrated the length of the control gate, but which were significantly

attenuated in the treatment gate. For example, on day 339, the flux of TOL at the end of

the control gate was 13 mg/m2/day compared to only 2.8 mg/m2/day in the treatment gate.

PCE underwent degradation in both gates, as indicated by the presence of its daughter

products, TCE and cDCE. However, in the control, both TCE and cDCE were only

detected in trace quantities ( < 30 Ag/l), and the extent of degradation process was too

limited to quantify (Devlin et al., 2002). In contrast, PCE was found to degrade

anaerobically in the treatment gate with a half-life of about 60 days, and there were

Fig. 12. Snapshots of the control (natural attenuation) and treatment gates on day 339. Maximum concentrations

at each monitoring well are contoured.


indications that the degradation rate had increased with time (based on ever increasing

levels of cDCE at G315). A mass balance calculation, and supporting flux calculations,

indicated that TOL also degraded in the control, with a half-life of about 60 days (Devlin

et al., 2002). In the treatment gate, TOL was found to degrade aerobically with a half-life

of about 20 days. These results represent notable improvements in TOL and PCE

degradation compared with natural attentuation.

Two degradation products of PCE were identified in the gates: TCE and cDCE. The

former compound never exceeded 15 Ag/l in either gate, and the latter compound never

exceeded 30 Ag/l in the control gate. However, cDCE was detected at 289 Ag/l at one point(G312M1) in the treatment gate, and the fence-averaged concentrations were increasing in

the anaerobic treatment zone, at G315 (Fig. 9), at the end of the experiment. Despite the

build-up of cDCE in the anaerobic part of the treatment gate, there was significant

degradation of the substance in the aerobic section, as planned. The result was that very

little cDCE arrived at the end of the gate, and indications were that the degradation rate for

this compound was increasing at the time the experiment ended.

Despite the promising outcome of this experiment, additional work is needed to

improve the efficiency of removal of PCE and TOL. To illustrate this, it can be shown

that for 99% PCE removal within the treatment gate (between G306 and G315), a

degradation half-life of about 8.5 days would be required, and for TOL to similarly

degrade (between G316 and G324) would require a half-life of about 8 days. The


estimated half-lives here were about 60 and 20 days, respectively. It should be noted that

the PCE degradation rate estimate may be conservatively low, since no PCE advance was

noted in the treatment gate between days 238 and 339, an interval over which there should

have been a clear invasion of the ROB by PCE. Also, TOL degradation may have been

more rapid if benzoate inputs had been scaled back. It should also be noted when the

experiment ended, there were indications that PCE and cDCE degradation rates were

increasing, so the goals of 8 day half-lives may yet be attainable.

5.8. Anaerobic–aerobic boundary

By sequencing anaerobic and aerobic treatment systems, possible interferences might

be expected. In particular, chemical precipitates, such as iron oxides, might be expected to

accumulate at the transition boundary. There is also the possibility of biofouling at the

boundary, due to the accumulation of aerobic bacterial biomass in response to the chemical

disequilibrium that exists there. If pore clogging occurred and compromised the treatment

Fig. 13. Hydraulic gradients through the treatment gate at the beginning and end of the experiment. The

consistently linear gradient indicates that no hydraulically significant fouling occurred in the anaerobic–aerobic

transition zone over the duration of the experiment.


gate performance, it would most likely manifest itself as a hydraulic barrier. In a passive

system (no pumping well) such a barrier would divert water around the gate. In the system

investigated here, where a pumping well ensured flow through the gate, the effect of the

barrier would be to create an hydraulic head drop across the redox transition zone.

Hydraulic head measurements were made routinely during the experiment (Fig. 13).

The hydraulic gradients were generally quite reproducible and the trends in hydraulic head

along the treatment gate were remarkably consistent over the entire 339 days of the

experiment. There is no indication in these data that clogged pores at the anaerobic–

aerobic boundary caused flow problems. Nevertheless, these data represent coarse

measurements made over a short duration experiment (f 1 year), and so they must be

considered preliminary. Additional work is planned to investigate this potential problem at

the pore scale. Further work is also needed to assess the longer term effects on

performance.

6. Conclusions

Based on the work presented above, it may be concluded that sequenced bioremedi-

ation is an advantageous and technically viable method of treating mixed contaminant

plumes. The validity of this conclusion is particularly well illustrated by the formation of

cDCE in the anaerobic zone, followed by its degradation in the aerobic zone. However, it

is also significant that the benzoate additions, with nutrient supplements, stimulated the

degradation of PCE, and TCE to form cDCE at all. Previous work in the same aquifer

failed to achieve chlorinated ethene transformations.

The treatment of mixed contaminants was successfully demonstrated with TOL. TOL

was found to degrade with a half-life of about 60 days by anaerobic natural attenuation

(Devlin et al., 2002). This process appeared to be inhibited in the anaerobic environment

created here. However, TOL degradation rates were tripled, compared to natural attenu-

ation, in the aerobic treatment zone.

Brief daily sparging for the emplacement of residual oxygen bubbles had no significant

effect on the aqueous concentrations of TOL or cDCE. The same conclusion is expected to

apply to other volatile compounds with Henry’s constants of similar magnitude. However,

this amount of sparging appears to be sufficient to create and sustain an aerobic

environment in which the target compounds can be degraded. Further effort is needed

to properly balance the supply of carbon in the NIW section with the oxygen demand in

the ROB section.

There was little advantage of the sequenced treatment over natural attenuation for the

chlorinated methanes. Both CT and CF were completely attenuated over about the same

distances in the treatment and control gates. These transformations occurred upstream of

the benzoate additions in the NIW, and were also observed in the control gate where

benzoate was altogether absent. This result suggests that TOL was serving as an electron

donor for the degradation of CT and possibly CF. However, it was noted that the lag time

before CF degradation began was several weeks shorter in the treatment gate than the

control, suggesting that the added benzoate was the dominant electron acceptor where it

was available.


Both TOL and PCE degraded more rapidly in the treatment gate than by natural

attenuation. Also, in both cases, the degradation rates were apparently increasing when the

experiment ended, so the maximum rates have yet to be determined.

Finally, there was no evidence of hydraulic interferences due to pore clogging, despite a

sharp transition form highly anaerobic conditions to aerobic conditions. Further work is

needed to examine the longer term performance of these coupled technologies and the

transition zone between them.

Acknowledgements

Financial support for this research was provided by the Advanced Alternative

Technology Demonstration Facility (AATDF), funded by the Department of Defense and

administered by Rice University. Additional support was provided by the NSERC/

Motorola/ETI Industrial Research Chair in Groundwater Remediation, and the Solvents in

Groundwater Consortium, University of Waterloo. T. Anderson of the University of

Massachusetts assisted with hydrogen measurements. The authors gratefully acknowledge

the comments, discussions and technical support given by M. McMaster, S. Froud, J.

Rosen, B. Butler, R.W. Gillham and S. Fiorenza.

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