Enhancement of in situ biodegradation of organic compounds in groundwater by targeted pump and treat intervention

Applied Geochemistry 48 (2014) 28–40

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Enhancement of in situ biodegradation of organic compounds in groundwater by targeted pump and treat intervention S.F. Thornton a,⇑, K.M. Baker b, S.H. Bottrell b, S.A. Rolfe c, P. McNamee b, F. Forrest b, P. Duffield b, R.D. Wilson a, A.W. Fairburn a, L.A. Cieslak a a b c

Groundwater Protection and Restoration Group, Kroto Research Institute, University of Sheffield, Sheffield S3 7HQ, United Kingdom School of Earth and the Environment, University of Leeds, Leeds LS2 9JT, United Kingdom Department of Animal and Plant Sciences, University of Sheffield, Sheffield S10 2TN, United Kingdom

a r t i c l e

i n f o

Article history: Available online 4 July 2014 Editorial handling by M. Kersten

a b s t r a c t This study demonstrates the value of targeted pump and treatment (PAT) to enhance the in situ biodegradation of organic contaminants in groundwater for improved restoration. The approach is illustrated for a plume of phenolic compounds in a sandstone aquifer, where PAT is used for hydraulic containment and removal of dissolved phase contaminants from specific depth intervals. Time-series analysis of the plume hydrochemistry and stable isotope composition of dissolved species (d34S-SO4, d13C-CH4, d13C-TDIC (TDIC = Total Dissolved Inorganic Carbon)) in groundwater samples from high-resolution multilevel samplers were used to deduce changes in the relative significance of biodegradation processes and microbial activity in the plume, induced by the PAT system over 3 years. The PAT system has reduced the maximum contaminant concentrations (up to 6800 mg L1 total phenols) in the plume by 50% to 70% at different locations. This intervention has (i) stimulated in situ biodegradation in general, with an approximate doubling of contaminant turnover based on TDIC concentration, which has increased from <200 mg L1 to >350 mg L1, (ii) enhanced the activity of SO4-reducing microorganisms (marked by a declining SO4 concentration with corresponding increase in SO4-d34S to values >7–14‰V-CDT relative to background values of 1.9–6.5‰V-CDT), and (iii) where the TDIC increase is greatest, has changed TDIC-d13C from values of 10 to 15‰V-PDB to 20‰V-PDB. This indicates an increase in the relative importance of respiration processes (including denitrification and anaerobic methane oxidation, AMO) that yield 13C-depleted TDIC over fermentation and acetoclastic methanogenesis that yield 13C-enriched TDIC in the plume, leading to higher contaminant turnover. The plume fringe was found to be a zone of enhanced biodegradation by SO4-reduction and methanogenesis. Isotopically heavy methane compositions (up to 47.8‰V-PDB) and trends between d13C-TDIC and d13C-CH4 suggest that AMO occurs at the plume fringe where the contaminant concentrations have been reduced by the PAT system. Mass and isotope balances for inorganic carbon in the plume confirm the shift in spatial dominance of different biodegradation processes and significant increase in contribution of anaerobic respiration for contaminant biodegradation in zones targeted by the PAT system. The enhanced in situ biodegradation results from a reduction in organic contaminant concentrations in the plume to levels below those that formerly suppressed microbial activity, combined with increased supply of soluble electron acceptors (e.g. nitrate) into the plume by dispersion. An interruption of the PAT system and recovery of the dissolved organic contaminant concentrations towards former values highlights the dynamic nature of this enhancement on restoration and relatively rapid response of the aquifer microorganisms to changing conditions induced by the PAT system. In situ restoration using this combined engineered and passive approach has the potential to manage plumes of biodegradable contaminants over shorter timescales than would be possible using these methods independently. The application of PAT in this way strongly depends on the ability to ensure an adequate flux of dissolved electron acceptors into the plume by advection and dispersion, particularly in heterogeneous aquifers. Ó 2014 Elsevier Ltd. All rights reserved.

⇑ Corresponding author. Tel.: +44 (0)114 222 5744; fax: +44 114 222 5700. E-mail address: s.f.thornton@sheffield.ac.uk (S.F. Thornton). http://dx.doi.org/10.1016/j.apgeochem.2014.06.023 0883-2927/Ó 2014 Elsevier Ltd. All rights reserved.

S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

1. Introduction Phenols (e.g. phenol and cresols) and their derivatives occur naturally in the environment but have also been produced commercially for use in many products. Many industrial activities, such as wood preservation and tannery facilities, fossil fuel refining processes, manufactured gas plants, pesticide and pharmaceutical production also create phenolic compounds as a by-product of their respective processes (Pereira et al., 1983; Kumaran and Paruchuri, 1997; Squillace et al., 1999; Broholm and Arvin, 2000; van Schie and Young, 2000; Broholm and Arvin, 2001; Michalowicz and Duda, 2007; Al-Khalid and El-Naas, 2012). Phenols have important antimicrobial properties, which have medical uses, but can be harmful with the potential to cause acute health affects (Bruce et al., 1987; Tsutsui et al., 1997; Goddard and McCue, 2001; Michalowicz and Duda, 2007) and pose an environmental hazard (Al-Khalid and El-Naas, 2012; Lin et al., 2012). Consequently, the release of these organic chemicals into groundwater from industrial activities will usually require appropriate restoration strategies. Pump and treatment (PAT) is a long-established and widely-used engineered restoration technique for contaminated groundwater (McKinney, 1992; US EPA, 1996; Cohen et al., 1997; Matott et al., 2006; Champagne et al., 2012), in which contaminated groundwater is pumped from an aquifer for ex situ treatment by chemical and biological processes (Suthersan, 1999; Simon et al., 2002; Champagne et al., 2012). Other applications of PAT include the injection of chemical reagents to enhance removal of adsorbed and free-phase organic contaminants (Palmer and Fish, 1992; Suthersan, 1999). PAT is also implemented for hydraulic containment to control the migration of contaminated groundwater, preventing continued expansion of the contaminant zone or plume (McKinney, 1992; US EPA, 1996; Cohen et al., 1997; Suthersan, 1999; Matott et al., 2006). In this respect, PAT can be very effective for the management of contaminant plumes which are present at a scale or depth in aquifers that extend beyond the technical feasibililty or range of other engineered interventions. Biodegradation processes which occur naturally in contaminated groundwater have been used for many years to support risk-based restoration concepts such as natural attenuation (NA) and in situ bioremediation (Wiedemeier et al., 1999). Using aqueous and mineral electron acceptors in aquifers indigenous microorganisms can biotransform a wide range of organic chemicals in contaminated groundwater, via pathways which include aerobic respiration, denitrification, sulphate-reduction, methanogenesis and fermentation (Bradley, 2000; Major et al., 2002; Spence et al., 2005). However, the spatial and temporal distribution of these biodegradation processes and, in turn, the metabolic activity of the microorganisms responsible can be influenced significantly by the availability of suitable electron acceptors and nutrients, as well as specific effects related to the contaminant matrix (e.g. chemical composition, concentration of individual compounds and substrate bioavailability) (Broholm and Arvin, 2000; Spence et al., 2001a,b; Haack et al., 2004; Al-Khalid and El-Naas, 2012; Baker et al., 2012). These effects can consequently control the location, extent and relative rates of contaminant turnover in plumes (Schreiber and Bahr, 1999; Lerner et al., 2000; Thornton et al., 2001a,b; Kota et al., 2004; Wilson et al., 2004; Tuxen et al., 2006). The ability to modify the hydrochemical conditions in plumes, by alleviating limitations on microbial metabolism, may offer the opportunity to enhance the biodegradation of organic contaminants in situ. However, it is well known that increasing the microbial conversion capacity will not lead to higher biodegradation rates when mass transfer is a limiting factor (Boopathy, 2000). For example, it is usually necessary to increase the supply

29

or bioavailability of substrates required for microbial growth and activity to enhance in situ biodegradation. For groundwater plumes this has often been achieved by introducing soluble electron acceptors, electron donors and nutrients into contaminated aquifers to stimulate the activity of indigenous microorganisms, either by direct injection of these amendments or passive addition using technologies such as permeable reactive barriers (Semprini et al., 1990; Bedient and Rifai, 1992; Cunningham et al., 2001; Major et al., 2002; Zhang et al., 2009; Lin et al., 2012). Combining engineered and passive technologies for groundwater restoration is a novel but under-utilised management approach (e.g. Haas, 1997; Dettmers et al., 2006). It offers the potential for more complete treatment of contaminated groundwater in situ and achievement of restoration objectives in timeframes which may not be possible using one technique. The integration of PAT with NA and bioremediation for the management of contaminated groundwater has seldom been considered in practice, since these approaches have a different technical objective. PAT is typically used to simply remove contaminated groundwater from aquifers, whereas NA and bioremediation are used to treat contaminants in situ. In the research described in this study, PAT has been implemented in a targeted manner at field-scale to remove organic contaminants from specific depths in the aquifer, as a basis to enhance the in situ treatment of the residual contaminants by NA. It can be envisaged that perturbation of the plume by the PAT system in this way could also negatively affect the activity and function of indigenous microbial populations which have adapted to the in situ conditions. Potential effects include the mixing of groundwater with different redox status (aerobic vs anaerobic), contaminant concentration and pH, which may result in substrate starvation and suppression of metabolic function of specific microorganisms (Harrison and Loveless, 1971; Karthikeyan et al., 2001; Chakraborty et al., 2010). This is undesirable if such changes reduce biodegradation rates, lower contaminant turnover and decrease treatment efficiency. Moreover, the timescale for adaptation of the in situ microbial community to the imposed conditions and re-establishment of optimum metabolic activity may limit the overall performance of the engineered restoration, if this occurs over a longer period than the amendment (e.g. the rate of groundwater abstraction in the case of PAT). In this paper we investigate the contribution of targeted PAT to enhance the in situ biodegradation of organic contaminants in groundwater for improved restoration performance. This approach is examined using a PAT system installed for hydraulic containment and treatment of a plume of phenolic compounds in a sandstone aquifer. The spatial distribution of microbially-mediated terminal electron accepting processes (TEAPs) and microbiological activity in this plume is influenced significantly by the availability of dissolved electron acceptors and concentration of the phenols (Lerner et al., 2000; Pickup et al., 2001). Biodegradation of the phenols is sensitive to the organic contaminant concentration and suppressed at certain concentrations (Harrison et al., 2001; Wu et al., 2002, 2006). Specifically, aerobic respiration and denitrification are restricted to a narrow zone at the plume fringe, SO4-reduction is suppressed within the plume at a contaminant concentration above 2000 mg L1 and fermentation processes are more important for mass turnover at higher contaminant concentrations (Thornton et al., 2001a,b; Spence et al., 2001b; Wu et al., 2002, 2006; Baker et al., 2012). This conceptual model of contaminant biodegradation potential in the plume has remained consistent since 2001. It informed the design of the PAT system, which has been operated to remove groundwater from specific depths in the plume, where biodegradation is inhibited by the contaminant concentrations. This conceptual model can thus be tested by examining the change in the distribution of TEAPs, caused by the PAT

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

system. The hypothesis evaluated is that biodegradation of the organic contaminants will be enhanced at the specific locations in the plume targeted by the PAT. This is expected to occur by (i) reduction of the contaminant concentrations to values below those that suppress the activity of microorganisms responsible for the TEAPs, and (ii) increased supply of dissolved electron acceptors for biodegradation, via mixing by dispersion at the plume-uncontaminated groundwater interface (plume fringe). This hypothesis is examined using groundwater samples collected in three separate years from high-resolution multilevel samplers (MLS) installed at two locations in the plume, close to abstraction wells used for the PAT system. The sampling period brackets the installation and operation of the PAT system at the site. Stable isotope analysis of dissolved sulphate, inorganic carbon and methane in the groundwater samples is used to interpret the spatial and temporal evolution of biodegradation processes over this period, from which the influence of the PAT system on in situ microbial activity in the plume is inferred. This approach has been used in other studies to establish the relative contribution of respiration processes and methanogenesis, and to elucidate the relevant controls on their distribution in groundwater (Hunkeler et al., 1999; Spence et al., 2001b; Baker et al., 2012). The theory underpinning the application of stable isotope analysis in environmental systems is well understood (see Hunkeler et al., 1999; Ahad et al., 2000; Meckenstock et al., 2004; Knöller et al., 2008 for review and discussion) and will not be elaborated here. To our knowledge this is the first field-scale study of enhanced in situ biodegradation using this type of engineered restoration in this ‘‘targeted’’ way. The specific objectives of this study were: 1. To identify changes in the organic and inorganic hydrochemistry of the plume after implementation of the PAT system using high-resolution multilevel groundwater samplers. 2. To deduce the relationship between dissolved organic contaminant concentrations in groundwater and the spatial and temporal distribution of in situ biodegradation processes and carbon cycling in the plume, in response to the operation of the PAT system. 3. To establish the contribution of the targeted PAT system for enhanced biodegradation of the organic contaminants in situ, by comparing the spatial distribution of specific microbial processes in the plume before and after implementation of the system. 2. Study site This study was conducted at a site on the Permo-Triassic Sherwood Sandstone aquifer in the UK. The aquifer is formed from fluvial and deltaic sequences of red and brown sands, gravels and conglomerates with interbedded marls (Allen et al., 1997; TylerWhittle et al., 2002). At the site these strata dip 2 degrees to the NW and lie beneath a thin cover of permeable glacial drift deposits (Williams et al., 2001). Groundwater beneath the site has become contaminated with phenolic compounds and mineral (sulphuric) acid during the operation of a former coal tar distillation plant (Thornton et al., 2001a; Williams et al., 2001). Site history and groundwater flow directions suggest that former contaminant releases have been similar in composition and that the plume flow direction is consistent, with a groundwater velocity that varies from 4 to 11 m yr1 depending on depth (Williams et al., 2001). The plume was delineated using multiple single screen observation wells before the installation of two high-resolution multi-level sampling (MLS) boreholes (MLS59 and MLS60) along the plume flow path (Fig. 1). The plume is deeper in the aquifer at the location of MLS60 compared with MLS59. This is attributed to the higher density of the contaminated groundwater relative to the

background groundwater, which causes the plume to ‘‘dive’’ at a shallow angle (Thornton et al., 2001a). The plume hydrochemistry and microbiology have been described in detail elsewhere (Pickup et al., 2001; Spence et al., 2001b; Thornton et al., 2001a; Williams et al., 2001; Rizoulis et al., 2012). The uncontaminated groundwater is aerobic. The water table fluctuates between 5 and 6 mbgl over an annual cycle, as deduced from the resolution of sampling ports on the MLS installations. A pump and treat (PAT) remediation system was implemented in 2009 for hydraulic containment of the plume and removal of contaminated groundwater for ex situ treatment in an onsite facility. Abstraction wells were installed at locations (P1 and P2) along the plume flow path, approximately 5 m transverse to MLS59 and MLS60, respectively (Fig. 1). The abstraction wells are partially screened and pumps were installed to extract groundwater from specific intervals (ca. 28 mbgl at P1 and 39 mbgl at P2) in the aquifer, corresponding approximately to the highest dissolved organic contaminant concentrations in the plume, supported by results obtained from the adjacent MLS (Thornton et al., 2001a) and groundwater flow modelling. A summary of the pumping and groundwater sampling schedule is given in Table 1. In general the pumping rate has increased over the monitoring period. The close proximity of the abstraction wells to the MLS boreholes allowed temporal changes in the plume chemistry during the operation of the PAT to be examined at high (1 m) resolution over a monitored depth up to 45 m (max depth in MLS60), by collection of groundwater samples from each MLS at approximately annual intervals from 2009 to 2012. The PAT system has been operated continuously from September 2009 at the location of P2 (adjacent to MLS60) and from November 2009 until July 2011 at the location of P1 (adjacent to MLS59), when it was stopped for maintenance. The yearly-averaged phenol concentration (mg L1) in groundwater extracted from pumping stations P1 and P2 during operation in 2009, 2010 and 2012 was 4944 ± 536, 5208 ± 837, 5374 ± 428 and 3178 ± 259, 2961 ± 322, 2324 ± 55, respectively (Mr. R Astbury, personal communication).

3. Materials and methods 3.1. Groundwater sampling and analysis Groundwater samples were collected in 2009, 2011 and 2012 from MLS59 and MLS60 according to the schedule in Table 1, using methods described in Thornton et al. (2001a). The MLS are instrumented with 5 cm screens spaced at metre intervals over the monitored depth; further details on their construction are provided in Thornton et al. (2001a). Samples were filtered (0.45 lm WhatmanÒ nylon) into 14 ml glass vials for the analysis of the phenolic compounds and into 15 ml polypropylene tubes for dissolved major ion and acetate analysis. Unfiltered samples were collected and dispensed into: 1L amber glass bottles containing 1 M (final concentration) CuCl2 and 15% v/v acetic acid (to precipitate dissolved sulphide and inhibit microbial activity) for d34S-SO4 analysis and 500 ml polypropylene bottles containing 2.5 M SrCl2 and 0.1 M NH4OH (final concentration) for the analysis of total dissolved inorganic carbon (TDIC) and d13C-TDIC analysis (Spence et al., 2001b; Baker et al., 2012). All vessels were pre-filled with oxygen-free N2, filled completely with sample to minimise atmospheric exposure and stored at 4 °C until analysis within 1 week of collection. Measurements were made for dissolved oxygen, pH and dissolved iron in the groundwater samples, according to Thornton et al. (2001a), but are not presented. The dissolved oxygen concentration varied from 4.1 to 10.4 mg L1 above the plume at MLS59 and from 1.8 to 2.1 mg L1 at MLS60 but was below detection in the plume. The pH varied from 6.5 to 8.1 at MLS59 and from 4.6 to 6.8 at MLS60

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

MLS borehole

~130m

Observation borehole

Pumping well

60

59

P2

P1

Site

Groundwater Flow (4 to 11 m/yr)

Fig. 1. Plan view of site showing location of multilevel sampling boreholes (MLS) and abstraction wells (P1 and P2) used for the PAT system. The dashed line denotes the approximate location of the plume envelope (modified from Thornton et al., 2001a).

Table 1 Pumping regime and groundwater sampling schedule during operation of pump and treat system. Pumping station P1

a b

Pumping station P2

Pumping period

Pumping rate (m3 h1)

Groundwater sampling date

Pumping period

Pumping rate (m3 h1)

Groundwater sampling date

November–December 2009 February–December 2010 January–June 2011b –

0.26–0.66

June 2009a

0.76–1.24

June 2009a

0.27–1.45 1.24–2.05 No pumping

No sampling June 2011 June 2012

September–December 2009 January–December 2010 January–June 2011b July 2011–July 2013

0.33–2.0 1.05–2.62 0.12–2.34

No sampling June 2011 June 2012

Initial groundwater quality survey undertaken before start-up of PAT system. Dates provided for equivalent comparison of pumping period between the stations.

(values were lower in groundwater above the plume at this location). The dissolved iron in the plume was the ferrous (Fe2+) form, according to the colorimetric analysis performed. Samples for dissolved methane were obtained using a gas stripping cell (Microseeps™) attached to the groundwater sampling pump at the well head. The method is analogous to the ‘‘bubble strip’’ technique used in similar studies (Lovley et al., 1994; Chapelle, 1997). A 20 ml volume of N2 was injected into the groundwater-filled cell and gas samples were removed after equilibration for 30 min under a constant sample stream (100 ml min1). Duplicate gas samples (10–15 ml) were injected into 20 ml N2-filled septum-sealed vials and analysed within 24 h by standard methods. Total phenols (phenol, o-cresol, m/p-cresol, dimethylphenol isomers) were analysed by high-pressure liquid chromatography, using a Perkin-Elmer Series 200 system and UV detector, with appropriate organic compound calibration standards and analytical quality controls. The method detection limit and precision was 1 mg l1 and ±5%, respectively. Major ions (including acetate, NO3, NO2 and SO4) were determined by ion chromatography using a Dionex 2000 system with a detection limit of 1 mg l1 and precision of ±3%, as described by Thornton et al. (2001a). Methane in gas samples was measured using a Perkin Elmer Clarus gas chromatograph calibrated with certified commercial standards. The method detection limits and precision were 10 ppm (v/v), ±2%. Trip sample blanks (mixed laboratory standards) were also analysed to identify changes in gas composition during field transportation and storage; no changes were detected in the standards. The headspace gas concentration and Henry’s Law constant for the groundwater temperature (12 °C) were used to obtain the dissolved CH4 concentration in groundwater samples. Estimates of the aquifer hydraulic conductivity were made using the MLS installations, to compare the spatial variability in this property with observed changes in the plume chemistry. Falling head tests were performed using the MLS ports on MLS59 and MLS60, by recording the time for a column of water to fall over a measured scale attached to the MLS tubing. The results were analysed using the Hvorslev method (Hvorslev, 1951; Sara, 2003). The precision of the estimates is ±5.3%, based on replicate tests.

3.2. Stable isotope analyses Sulphate was recovered from groundwater samples by precipitation with BaCl2 at pH 2.5–3 followed by filtration. The recovered BaSO4 precipitate was quantitatively converted to SO2 by combustion at 1150 °C in the presence of pure oxygen (N5.0) injected into a stream of helium (CP grade). N2 continued through the system unchecked whereas SO2 was removed from, and re-injected into, the gas stream using temperature controlled adsorption/desorption columns. The d34S (reported in parts per thousand ‰ relative to the IAEA Vienna-Cañon Diablo Troilite (V-CDT) standard) is derived using the integrated mass 64 and 66 signals relative to those in a pulse of SO2 reference gas (N3.0) and calibrated to the international V-CDT scale using an internal laboratory BaSO4 standard derived from seawater (SWS-3). This has been analysed against the international standards NBS-127 (+20.3‰), NBS-123 (+17.01‰), IAEA S-1 (0.30‰) and IAEA S-3 (32.06‰) and assigned a value of +20.3‰, and an inter-lab chalcopyrite standard CP-1 assigned a value of 4.56‰. If samples are more 34S depleted than CP-1, the IAEA S-3 standard is used instead. The SrCO3 precipitated in the 500 ml sample bottles was carefully collected on a pre-weighed filter, washed to remove reagent residues, oven-dried (90 °C for 24 h) and reweighed to determine the recovery of SrCO3. This precipitate was reacted with phosphoric acid under vacuum to liberate CO2, which was cryogenically purified and determined manometrically, then used to calculate TDIC from the mass of SrCO3 recovered from the groundwater. The precision of this analysis varies with TDIC concentration, but is better than 2% at all concentrations encountered in this study; blank contributions (analysed from reagents in unfilled sample bottles) were insignificant, always less than 0.01% of sample yields. Methane samples were cryogenically purified before combustion to CO2 using CuO and a Pt catalyst on a purpose-built vacuum line. Carbon isotope analysis was performed on CO2 gas using a VG SIRA-10 gas source mass spectrometer. Data were corrected using standard procedures (Craig, 1957) and results reported as d13C in ‰ relative to Vienna-PeeDee Belemnite (V-PDB). The precision of the isotopic analyses is 0.3‰ for d34S and methane-d13C, and better than 0.1‰ for TDIC-d13C.

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40 Table 2 Composition of uncontaminated groundwater sampled at MLS59 and MLS60 between 2009 and 2012.

4. Results and discussion 4.1. Spatial variability in aquifer hydraulic conductivity

Species

The hydraulic tests undertaken on the MLS installations and also reported previously for other boreholes across the site (Thornton et al., 2001a) show that the variation (0.72– 3.23 m day1) in aquifer hydraulic conductivity (K) is within the range considered suitable for enhanced bioremediation (Bedient and Rifai, 1992; Cohen et al., 1997) and similar to published values for this aquifer formation (Allen et al., 1997). It is higher at shallow depth but decreases by only a factor of two over the plume section and is very similar at both locations (Supplementary Fig. 1). The overall variance in ln K (0.13) is much lower than reported for many other aquifers (Gungor-Demirci and Aksoy, 2011). The aquifer formation has a relatively low mean anisotropy ratio (ratio of horizontal hydraulic conductivity to vertical hydraulic conductivity) of 1.1 (Allen et al., 1997). This feature and the absence of significant physical heterogeneity in the aquifer means that the PAT system is therefore likely to be very effective in removing contaminant mass and, in turn, directly modifying the plume geometry and chemistry under the applied flow field (MacKay and Cherry, 1989; Lee et al., 2000; Ko et al., 2005). 4.2. Influence of pumping on the spatial and temporal variation in contaminant concentration Concentration-depth profiles of total phenols concentration, TPC (sum of phenol, cresols and dimethylphenols) are presented in Fig. 2. Depth (m) is given in metres below ground level (mbgl) and the water table is 5 mbgl at both locations. The TPC distribution shows that the plume remained within the depth intervals (10–30 mbgl at MLS59 and 20–45 mbgl at MLS60) reported by Thornton et al. (2001a) and Baker et al. (2012) during the monitoring period. Uncontaminated groundwater sampled above the plume at MLS59 (6–10 mbgl) and MLS60 (6–20 mbgl) represents the background groundwater chemistry at these locations, based on the results in Thornton et al. (2001a) and repeated sampling since then. A summary of the major ion composition of groundwater sampled from these intervals between 2009 and 2012 is given in Table 2. This is very similar to published compositions of uncontaminated groundwater in this aquifer formation (Tyler-Whittle et al., 2002).

Cl NO 3 SO2 4 Ca2+ Mg2+ Na+ K+ Mn2+ Fe2+

MLS59a 2009

2011

2012

2009

2011

2012

28.3 62.8 27.0 108.6 7.4 26.2 4.5 0.01 0.03

12.9 67.7 25.1 92.7 9.3 15.2 1.2 0.01 0.4

14.2 48.8 27.6 80.1 8.2 15.1 1.5 0.01 0.21

41.5 62.1 42.9 67.3 10.2 13.9 5.6 0.09 BDLb

22.0 87.7 35.1 53.1 11.3 9.5 4.5 0.07 0.23

20.4 57.8 36.4 56.6 12.9 7.8 5.6 0.1 0.09

a No organic chemicals were measured above detection limits on any survey and the inorganic chemical composition is presented for selected species only. b BDL: Below method detection limit.

The PAT system was installed to remove contaminated groundwater from depths with the highest concentration of phenolic compounds in the monitored profile and therefore reduce contaminant flux. A comparison of the TPC depth profiles for MLS59 and MLS60 between 1998 and 2012 shows that the contaminant distributions in the plume prior to the installation of the PAT system are relatively consistent (Supplementary Fig. 2). It also highlights the significant and rapid influence of the PAT operation on TPC distribution and concentration within the specific depth intervals targeted after system start-up in 2009. This behaviour was captured by the groundwater sampling from 2009 to 2012, which brackets continuous operation of the system at MLS60 and operation at MLS59 until 2011 (Fig. 2). The PAT system shutdown after June 2011 allowed the re-establishment of ambient groundwater flow conditions at the location of MLS59. Between 2009 and 2011 the TPC between 24 and 28 mbgl at MLS59 was decreased from 4600 mg L1 to below 1500 mg L1 (70% reduction) by the PAT system. At the same time breakthrough of nitrate into the plume occurred, to a maximum of 20 mg L1 from 27 to 30 mbgl (Fig. 3a). The nitrate concentration in the plume remained below detection on previous surveys as consumption by denitrification outcompeted inward migration of nitrate from the uncontaminated groundwater (Spence et al., 2001a; Thornton et al., 2001a). The detection of nitrate in the plume under these new conditions results from increased mixing

TPC (mg L-1)

a

0

1000

2000

3000

TPC (mg L-1) 4000

5000

b

0 5

25

2000

4000

6000

8000

5

2009 2011 2012

10 15

Depth (mbgl)

Depth (mbgl)

20

0 0

10 15

MLS60a

20 25 30 35 40

30 45 35

50

Fig. 2. Depth profiles of TPC in groundwater at the location of (a) MLS59 and (b) MLS60 before (2009) and after (2011 and 2012) the implementation of the groundwater pump and treatment system. The legend applies to both graphs. The water table is approximately 5 mbgl at both locations.

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

NO3 (mg L-1)

a

0

20

40

NO3 (mg L-1)

60

80

100

0

b

0

20

40

60

80

100

120

0 5

5

10

Depth (mbgl)

Depth (mbgl)

10 15 20 25

2009 2011 2012

30

20 25 30 35 40 45

35

50 -1

SO4 (mg L )

c

15

0

50

100

150 0

δ34S (‰)

5

10

15

SO4 (mg L-1) 20

0

d

0

200

400

0

δ34S (‰) 5

10

15

0 5

5

10

Depth (mbgl)

Depth (mbgl)

10 15 20 25

15 20 25 30 35 40

30

45

35

50

TDIC (mg L-1)

e

0

TDIC (mg L-1)

δ13C (‰)

100 200 300 400 -24 -22 -20 -18 -16 -14

0

f

0

200

400

δ13C (‰) -30 -25 -20 -15 -10 -5

0

0 5

5

10

Depth (mbgl)

Depth (mbgl)

10 15 20 25

15 20 25 30 35 40

30

45

35

50

CH4/(CH4+CO2)

g

0.00 0.05 0.10 0.15 0.20 -60

CH4(CH4+CO2)

13

δ C (‰) -55

-50

-45

0

h

0.00

0.01

0.02

13

δ C (‰) 0.03-60 -55 -50 -45 -40 -35 -30

0 5

5

Depth (mbgl)

10 15 20 25

Depth (mbgl)

10 15 20 25 30 35 40 30 35

45 50

Fig. 3. Depth profiles of concentration and isotopic composition of chemical species in groundwater at the location of MLS59 (left panels) and MLS60 (right panels) between 2009 and 2012. The legend applies to all graphs and the following features can be identified at both MLS locations: BG: Background groundwater; PF: Plume fringe (upper and lower definition); P: Plume. The water table is approximately 5 mbgl at both locations.

of uncontaminated groundwater into the plume due to the pumping, or the presence of converging flow paths en route to the MLS sampling port. This occurs because the capture zone of the

abstraction wells during pumping is 150 m (R Astbury, personal communication). Hence, the abstraction wells induce mixing between the plume and uncontaminated groundwater transverse

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

to the plume flow path during pumping (see locations in Fig. 1). The effect of the pumping is to increase the flux of nitrate into the plume for biodegradation. However, as this is an open system and the nitrate results are from single surveys rather than timeintegrated, it is not possible to establish a mass balance for this flux. After cessation of pumping in 2011 at the location of MLS59, the TPC in the plume increased to above 3000 mg L1 at 25 and 26 mbgl, with a corresponding decrease in the nitrate concentration. This records recovery of the TPC at this time towards the pre-pumping situation and renewed influence on the plume organic chemistry of the source term observed prior to the pumping (Baker et al., 2012). This recovery is commonly observed after the cessation of a PAT system, due to the transport of dissolved contaminants from upstream of the pumping location or contaminant desorption from the aquifer sediment (Hall and Johnson, 1992; Cohen et al., 1997; Simon et al., 2002). As there is limited attenuation of the phenols by sorption in this aquifer (Thornton et al., 2001a), the change in TPC at MLS59 after cessation of pumping is more likely to reflect the migration of dissolved contaminants from upstream of this location. The continuous pumping at the location of MLS60 has resulting in a significant decrease in the maximum plume TPC, from 6800 mg L1 in 2009 to 3400 mg L1 in 2011. The reduction in TPC was most pronounced between 30 and 45 mbgl, where it was up to 50% less than the pre-pumping value. This trend was evident in 2012 for all depths from 34 mbgl to the base of the monitored profile; the TPC at this time was the lowest measured since 1998 (Thornton et al., 2001a) (Supplementary Fig. 2b). As with MLS59, nitrate was measured inside the TPC plume at MLS60 for the first time since investigations began (Fig. 3b). The pump at MLS60 is situated at approximately 39 mbgl, and so the pumping has a lesser effect on flow paths towards the upper part of the plume (28–33 mbgl), where the maximum TPC exceeded 6000 mg L1 in 2012, comparable to previous surveys (Baker et al., 2012). The TPC at this shallow depth is influenced more by variation in contaminant flux from the source, as shown by Thornton et al. (2001a). Ultimately the maximum TPC in the plume at the location of MLS59 and MLS60 during pumping was the lowest over the monitoring period from 1998 (Supplementary Fig. 2; Thornton et al., 2001a; Baker et al., 2012). This has imposed important changes on the in situ hydrochemical conditions supporting contaminant biodegradation. 4.3. Spatial and temporal variability in indicators of in situ biodegradation processes The concentration of dissolved nitrate, sulphate and acetate, and the isotopic composition of dissolved sulphate (d34S-SO4), methane (d13C-CH4) and total dissolved inorganic carbon (TDIC, d13C-TDIC) measured in groundwater samples collected from MLS59 and MLS60 between 2009 and 2012 was used to evaluate the spatial and temporal variability of biodegradation processes in the plume after implementation of the PAT system. According to Thornton et al. (2001a) this suite characterises the contribution of fermentation (acetate and methane), respiration (nitrate and sulphate), inorganic biodegradation products (TDIC) and relative significance of these pathways (d34S-SO4, d13C-CH4 and d13C-TDIC) for biodegradation in the plume. Dissolved oxygen is not presented due to analytical interference from the contaminant matrix. However the distribution of aerobic respiration processes is confined to the plume fringe, where the nitrate profile is an effective proxy to describe dissolved electron acceptor inputs into the plume from the uncontaminated groundwater and sulphate is, moreover, present in the plume at a far higher concentration due to mineral acid releases at the site (Lerner et al., 2000; Thornton et al., 2001a).

At the upper plume fringe (10–12 mbgl) at MLS59 the nitrate concentration decreased markedly across the interface between the background groundwater and the plume (Fig. 3a). Very low values occur in the plume in 2009, but in 2011 and 2012 the nitrate concentration was higher (5–20 mg L1) at the lower plume fringe. At MLS60 the nitrate concentration within the plume was below detection in 2009, but up to 2.4 mg L1 in 2011 and 2012, with the highest value at 35 mbgl in 2011 (Fig. 3b). Also, the nitrate concentration reached 7 mg L1 at the lower plume fringe (44– 45 mbgl) in 2012. The sulphate concentration in the plume at MLS59 was much lower in 2011 than other years (Fig. 3c). It was isotopically heavy (8.9–15.2‰) at the upper plume fringe and from 12 to 14 mbgl, which coincided with a low sulphate concentration (Fig. 3c). This corresponds to a zone of established SO4-reduction in the plume at shallow depth, as deduced in previous studies (Spence et al., 2001b), which is consistent for the 2011 and 2012 surveys. The sulphate became isotopically lighter with depth and results from 2012 were comparable to 2011 except 25–27 mbgl, where sulphate was enriched in d34S (up to 13‰V-CDT) in 2011. However, sulphate was isotopically lighter (7‰V-CDT) at these depths in 2012. At MLS60 the sulphate concentration was lower within the plume in 2011 and 2012 than in 2009. The isotopic composition of dissolved sulphate was determined in 2011 and 2012 only. In 2011 the d34S composition of SO4 in the plume between 22 and 26 mbgl was isotopically heavy (up to 9.5‰V-CDT) compared with the background groundwater (1.9–6.5‰V-CDT). Below these depths (27–36 mbgl), d34S-SO4 was between 4.3 and 6.7‰V-CDT, but much heavier (up to 14‰V-CDT) from 36 to 45 mbgl. In 2012, d34S-SO4 values were comparable to those of 2011, except from 18 to 23 mbgl, where the sulphate was isotopically heavier (7.6–11.3‰V-CDT). The concentration of total dissolved inorganic carbon (TDIC) in groundwater above the plume at MLS59 was between 48 and 67 mg L1 in 2011, but generally higher within the plume (Fig. 3e) and at all sampled depths in 2012. The d13C-TDIC composition varied from 19.4 and 17.8‰V-PDB above the plume, but from 23.2 and 14.3‰V-PDB within the plume (Fig. 3e), with the lightest values between 26 and 29 mbgl (21.5 to 23.2‰V-PDB). A similar trend was observed in 2012, with the exception of isotopically heavier TDIC from 26 to 29 mbgl (17.4 to 19.5‰V-PDB). At MLS60 the TDIC concentration generally increased with depth and was higher at all depths in 2012 than 2011 (Fig. 3f). The TDIC was isotopically lighter from 29 to 45 mbgl in 2012 (18.1 to 22.9‰V-PDB), compared with 2011 (5.5‰ to 18.3‰V-PDB). The concentration and isotopic composition of methane in groundwater was determined in 2012 only. The methane concentration is presented as a proportion of the total dissolved gas measured in groundwater samples (Fig. 3g). Methane formed the highest proportion of dissolved gases at the upper fringe (12 mbgl) of the plume at MLS59. It decreased sharply over a 2 m interval (13 and 15 mbgl) and was then constant to 30 mbgl. The methane was also isotopically heavier (up to 47.8‰V-PDB) at the upper fringe (10–12 mbgl) and lower fringe (30 mbgl) of the plume, compared with the plume core (13–28 mbgl), where d13C-CH4 values as low as 56.4‰V-PDB were measured. At MLS60 dissolved methane was the highest proportion in the plume core at 26–35 mbgl (Fig. 3h). It decreased above and below these depths and was isotopically heavier in these zones (20 mbgl and 34–45 mbgl) than elsewhere. The concentration of acetate, an important fermentation product and substrate for microbial respiration, was 151 mg L1 at the upper fringe of the plume at MLS59, but decreased to 30 mg L1 at 18 mbgl (Fig. 4). At MLS60, the acetate concentration was higher (up to 140 mg L1) in the upper (22– 24 mbgl) and lower (42–45 mbgl) sections of the plume, compared with 7–19 mg L1 at 35–41 mbgl.

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

acetate (mg L-1) 0

a

50

100

acetate (mg L-1) 150

200

0

0

b

50

100

150

200

0 5

5

10

15 20 25

15

Depth (mbgl)

Depth (mbgl)

10

20 25 30 35 40

30 35

45 50

Fig. 4. Depth profiles of acetate concentration in groundwater at the location of (a) MLS59 and (b) MLS60 in 2012. Gaps in data at MLS59 are failed monitoring points. The absence of data from 25 to 34 mbgl at MLS60 is due to prioritisation of the groundwater sampling. The water table is approximately 5 mbgl at both locations.

4.4. Stimulation of natural attenuation by engineered remediation It was hypothesised that the relative significance of different biodegradation processes in the plume would be influenced by the change in the in situ hydrochemical conditions induced by the operation of the PAT system. Previous studies have demonstrated that the spatial distribution of biodegradation processes and activity of specific microbial communities in the plume are influenced significantly by the contaminant concentration (Harrison et al., 2001; Pickup et al., 2001; Spence et al., 2001b; Thornton et al., 2001a; Baker et al., 2012; Rizoulis et al., 2012). Specifically, anaerobic respiration processes appear to be suppressed at a TPC over 2000 mg/L (Harrison et al., 2001; Spence et al., 2001b; Wu et al., 2006). The mechanism responsible is unclear but this observation is consistent with the known antimicrobial properties of phenols (Rubbo and Gardner, 1965; Eiroa et al., 2005; Michalowicz and Duda, 2007; Greenberg et al., 2008). It probably results from a suppression of metabolic function (e.g. inhibition of enzyme synthesis or disruption of cell structure) in specific microorganisms by the contaminant matrix at elevated TPC (Wang et al., 1991; Eismann et al., 1997; Kumaran and Paruchuri, 1997; Chapman, 2003; Al-Khalid and El-Naas, 2012; Baker et al., 2012). Inhibition of microbial activity at high substrate concentration has been commonly observed in the biological treatment of industrial wastewaters containing phenols (Collins and Daugulis, 1997; Kumaran and Paruchuri, 1997; Goudar et al., 2000; Eiroa et al., 2005; Al-Khalid and El-Naas, 2012). It follows that the potential for contaminant biodegradation in the plume is also influenced by the distribution of TPC, if the respective respiration processes (e.g. aerobic respiration, denitrification, metal oxide-reduction, sulphate reduction) are affected by this variable, given their theoretical (stoichiometric) contribution to contaminant turnover. Hence, a reduction in the TPC to values that do not limit microbial activity should enhance biodegradation in situ. Conversely, this intervention may also adversely affect microbial processes which are adapted to the current plume conditions (Guieysse et al., 2001). This effect has been proposed (Semprini et al., 1990; Major et al., 2002), but not demonstrated at the field-scale. The TDIC concentration in groundwater increased in areas of the plume where the TPC was decreased after the PAT system started, for example between 35 and 44 mbgl at MLS60 (compare Figs. 2 and 3). This can be explained by increased TDIC production from biodegradation (via biogenic CO2) and shows the PAT system has enhanced the in situ microbial activity in general, with a

consequent increase in contaminant turnover. The isotopic fractionation studies therefore provide insight on which biodegradation processes have been stimulated to contribute to this increased TDIC, and the relative enhancement of these processes in the plume caused by the PAT system. Bacterial sulphate reduction (BSR) is an important biodegradation process in this aquifer and is suppressed by the contaminant matrix above a TPC of 2000 mg L1 (Pickup et al., 2001; Spence et al., 2001b; Baker et al., 2012). This TPC is within the range found to suppress the activity and growth of phenol-degrading microorganisms in laboratory studies using pure cultures (Goudar et al., 2000) and mixed cultures from this plume (Harrison et al. 2001; Wu et al., 2006) and other locations (Broholm and Arvin, 2000; Eiroa et al., 2005; Al-Khalid and El-Naas, 2012). As the TPC decreases below this value, the significance of BSR in the plume is therefore expected to increase. This feature was observed at MLS59 in 2011, when the TPC in the plume was reduced below 2000 mg L1 by the PAT system for the first time since monitoring began (Thornton et al., 2001a; Baker et al., 2012). The enhancement of in situ BSR is confirmed by the relative enrichment of sulphate 34S in the plume from 25 to 27 mbgl (with values > 7‰V-CDT, which exceed the d34S of the mineral acid released, Supplementary Table 1). BSR was negligible at these depths prior to the PAT system (Spence et al., 2001b; Baker et al., 2012) (Fig. 5a). Moreover, the TDIC-d13C was less than 21‰V-PDB at these depths in 2011 (Fig. 6a). This contrasts with more enriched d13C-TDIC values of approximately 15‰V-PDB in 2003 (Fig. 6b), when the TPC exceeded 6000 mg L1 and biodegradation was suppressed (with a consequent lower production of isotopically lighter biogenic TDIC). Hence the shift in the isotope composition of residual SO4 and TDIC confirms the stimulation of BSR in the plume by the PAT system in a zone where this microbial activity was previously suppressed under the ambient conditions. As nitrate was above detection in this part of the plume in 2011, biodegradation of the phenol compounds coupled to denitrification may also have contributed an isotopically light input to the TDIC pool (Guieysse et al., 2001; Harrison et al., 2001; Spence et al., 2001a). The importance of this biodegradation process at the plume fringe, due to mixing of uncontaminated groundwater into the plume by transverse dispersion, has previously been highlighted (Cirpka et al., 1999; Thornton et al., 2001a,b; Williams et al., 2001). It is generally accepted that the mixing of dissolved electron acceptors with organic contaminants by transverse dispersion is a key control on biodegradation rates at the plume fringe (Lerner et al., 2000; Pickup et al., 2001; Thornton et al., 2001a,b, 2011; Spence et al.,

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40 34

34

δ S-SO4 (‰) 02

a

4

6

8

10

δ S-SO4 (‰) 12

14

16

0

b

0

2

4

6

8

10

5

5

16

15

Depth (mbgl)

Depth (mbgl)

14

2003 2011 2012

10 10

12

0

15 20

20 25 30 35

25

40 30

45

35

50

34

Fig. 5. Comparison of d S-SO4 profiles from 2003, 2011 and 2012 at (a) MLS59 and (b) MLS60. The average and range of isotopic composition of the released mineral (sulphuric) acid (see Supplementary Table 1) are shown as dashed and dotted lines respectively. Positive excursions from this line provide evidence of BSR. The legend applies to both graphs. The water table is approximately 5 mbgl at both locations.

δ13C-TDIC (‰) -24

a

-22

-20

-18

-16

δ13C-TDIC (‰) -14

-12

0

-30

b

5

5

10

15 20 25

15

Depth (mbgl)

Depth (mbgl)

10

-25

-20

-15

-10

-5

0

0

2003 2011 2012

20 25 30 35 40

30 35

45 50

13

Fig. 6. Comparison of d C-TDIC profiles from 2003, 2011 and 2012 at (a) MLS59 and (b) MLS60. The water table is approximately 5 mbgl at both locations.

2005; Tuxen et al., 2006; Zhang et al., 2009). The transverse vertical dispersivity of this aquifer has been estimated to be 0.4 mm (Huang et al., 2000). Transverse horizontal dispersivity is typically higher than the vertical component (Fetter, 1993), but has not been estimated for the aquifer. Reactive transport modelling studies show that transverse vertical dispersion has an important influence on the hydraulic characteristics of the aquifer, by restricting vertical mixing of the plume (Mayer et al., 2001). This is also suggested by the hydraulic test data, which shows a relatively limited range in horizontal hydraulic conductivity over the vertical profile of the plume. It is likely that the PAT system has enhanced groundwater mixing at this interface, thereby stimulating contaminant biodegradation in the plume by increasing the supply of nitrate for denitrification. This can occur by the pumping increasing the contact area for transverse dispersion (Cirpka and Attinger, 2003; Dentz and Carrera, 2005). Modelling studies simulating the effect of hydraulic manipulation on biodegradation of this and other plumes have demonstrated the technical and practical feasibility of this enhancement (Marquis and Dineen, 1994; Jones et al., 2002; Zhang et al., 2009). At MLS60 BSR was also stimulated by the PAT system in 2011 and 2012. At this location there was relative enrichment of the plume SO4 in 34S at only at two depths on one sampling survey

between 1998 and 2003 (Spence et al., 2001b and Fig. 5b). However, in 2011 and 2012 the decrease in contaminant concentration due to the PAT system coincided with an enhancement in BSR in the upper part of the plume fringe (18–24 mbgl), with d34S values of up to 11.3‰V-CDT. This also occurred between 37 and 41 mbgl, marked by slight enrichment in 34S (up to 8‰V-CDT) over that of the mineral acid (7‰V-CDT) released from the plume source (Supplementary Table 1). In contrast, the relative significance of BSR in the plume at MLS59 was less in 2012 compared with 2011, following shut down of the PAT system. As each sampling survey is a single ‘‘snapshot’’ over the period indicated, the rate of change in BSR cannot be deduced. However, this change is important at the borehole-scale, when compared with the contaminant distribution. At this time (2012) the TPC increased to levels (>2000 mg L1) similar to those before system start-up in 2009. This was marked by a decrease in BSR in this part of the plume (25–27 mbgl). It is shown by d34S-SO4 values which are similar to those of the mineral acid released from the plume source (Fig. 5a) and heavier d13C-TDIC values (Fig. 6a), which imply a decrease in contaminant biodegradation at this time. Clearly, the operation of the PAT system has a pronounced and rapid impact (given the relatively slow plume migration) on the stimulation and relative significance of BSR in the plume. The

S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

aquifer microorganisms appear to respond relatively quickly to changes in plume chemistry induced by the PAT system and such changes can have both positive and adverse effects on the in situ microbial activity. This occurs primarily through (i) modification of the organic contaminant concentration (with attendant influence on microbial respiration in general and BSR in particular), and (ii) increased supply of dissolved oxidants in groundwater which support more energetically favourable anaerobic biodegradation processes (e.g. denitrification) in the plume. It would be interesting to determine at the field-scale if such changes are caused by the replacement (i.e. competition) of microbial populations or the adaptation of existing microorganisms (e.g. Guieysse et al., 2001), as has been examined in related laboratory process studies (Elliott et al., 2010). The concentration and isotopic composition of methane in the groundwater provides information on the relative contribution of methanogenesis and anaerobic methane oxidation (AMO) to TDIC production during and after operation of the PAT system (Fig. 6). AMO utilises isotopically light methane (leaving an isotopically heavy residual methane pool) and adds isotopically light carbon to the TDIC pool (Whiticar, 1999). A linear relationship should therefore exist between the isotopic composition of the two species during AMO; as d13C-TDIC becomes lighter, d13C-CH4 becomes heavier. The relationship between d13C-TDIC and d13C-CH4 from MLS60 (Fig. 7) shows a linear trend in 2012 at all depths except 28–33 mbgl (the interval where TPC > 4800 mg L1). The relatively heavy d13C-CH4 values above and below these depths at MLS60 confirm methane oxidation as a more important process at this time, suggesting that AMO occurs at the plume fringe where the TPC has been decreased by the PAT system (by either direct removal of contaminants or dilution with uncontaminated groundwater). Accordingly, AMO will have contributed to the observed increase in TDIC in the plume. The oxidation of methane under the higher redox potential found at the plume fringe is consistent with observations from other studies which show methane may be transported downstream and compete with pollutants for electron acceptors (Feisthauer et al., 2012). Where microbial respiration is suppressed by the TPC other anaerobic biodegradation processes and the activity of different

δ13 C-TDIC (‰) -30

-28

-26

-24

-22

-20

-18

-16

-10

δ13C CH4 (‰)

-20

MLS59 MLS60 MLS60 (High TPC)

-30

-40

37

microorganisms may contribute to organic contaminant turnover and accumulation of TDIC in the plume. Methanogenesis is a relatively important process in zones of high TPC in this plume (Baker et al., 2012). This most likely reflects the ability of Archaeal organisms present therein to tolerate the contaminant matrix (up to 10 g L1; De Rosa et al., 1986), presumably due to a more resistant cell membrane compared with bacteria. The tolerance of bacteria and other microorganisms to high concentrations of phenols has been attributed to specific modification of cell membrane physiology and composition (van Schie and Young, 2000). Methanogens can use acetate as an electron donor, producing isotopically light methane and isotopically heavy TDIC (Laukenmann et al., 2010). The light d13C-CH4 values (<-50‰V-PDB) from 28 to 33 mbgl at MLS60 show that methanogenesis is still relatively significant. This zone has the highest proportion of methane, relative to the total dissolved gases, which confirms the importance of this process. Thus, the PAT system has reduced the contaminant concentration at the location of MLS60 to an extent that microbial respiration processes, including BSR, AMO and (potentially) denitrification, have been stimulated and become spatially more important in the plume than prior to this intervention. This improvement will increase the overall contaminant turnover in the plume and potential for natural attenuation, where the TPC is below that which suppresses microbial respiration. In contrast, methanogenesis was still important at most depths in the plume at MLS59 in 2012, after the shutdown of the PAT system at this location and subsequent increase in the TPC. This is highlighted by much lighter d13C-CH4 values (approx. 55 to 50‰V-PDB), compared with MLS60, and absence of a strong relationship between d13C-CH4 and d13C-TDIC data (Fig. 7). However, the accumulation of TDIC at all depths at MLS59 between 2011 and 2012 suggests that respiration processes were spatially more significant until the cessation of pumping. This feature, with the isotopic evidence from the TDIC and sulphate plus nitrate breakthrough in 2011, illustrates the effect the PAT system has had in stimulating in situ microbial activity supporting different respiration processes and enhanced biodegradation of organic contaminants in the plume. To provide a semi-quantitative assessment of changes in microbially-mediated carbon turnover the origins of TDIC in the plume were ascribed to different processes by undertaking a mass and isotope balance for inorganic carbon (see e.g. Hunkeler et al., 1999). Since the TDIC concentration increased at all depths following the start-up of the PAT system, a simple mass balance model was used to estimate the d13C of the added TDIC, following the relationship:

d13 CFinal :½TDICFinal ¼ d13 CStart :½TDICStart þ d13 CAdded :½TDICAdded The origin of the added TDIC was estimated from its d13C composition, by apportioning the carbon to two sources: respiration CO2 (d13C = 26‰, equal to the isotopic composition of the organic contaminants, Spence et al. 2001b) and methanogenic CO2. The isotopic composition of methanogenic CO2 was estimated assuming that methane and CO2 were produced by acetoclasis from acetate with d13C = 26‰ (see Baker et al., 2012) and produced methane with the d13C value measured in the 2012 profiles. The carbon mass balance for the reaction then dictates that:

-50

2  ð26Þ ¼ 1  ðd13 CMethane Þ þ 1  ðd13 CMethanogenic CO2 Þ -60 Fig. 7. Relationship between d13C-TDIC and d13C-CH4 at the location of MLS59 and MLS60 in 2012, showing two distinct sets of values, which suggest some depth intervals at MLS60 represent similar conditions to the intervals at MLS59. The dashed line is a regression (with 95% confidence intervals) of the MLS60 isotope data where TPC were below 3600 mg L1.

The source of the initial TDIC has been estimated from its d13C composition using the same end-member compositions. The results are given in Table 3. The lower part of MLS60 shows the largest changes, with TDIC production prior to operation of the PAT system being dominated by methanogenic CO2 production (estimated at 89% of TDIC produced). The impact of the PAT system

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S.F. Thornton et al. / Applied Geochemistry 48 (2014) 28–40

Table 3 Contribution of respiration processes and methanogenesis to TDIC production in plume.

a b

MLS

Depth (mbgl)

TDIC increase (mg L1)

d13C of added TDIC (‰V-PDB)

59 59 59 59 60 60 60

13 16–22 25 >26 18–25 34–42 >42

200 200 200 200 290 230 85

18.6 18.0 19.6 17.2 20.1 26.8a 35.0b

% C from methanogenesis

% C from respiration

Start

Added

Start

Added

42 29 32 13 44 89 89

29 26 21 28 56 6a

58 71 68 87 25 11 11

71 74 79 72 75 106a

b

b

13

A very negative d C value of added TDIC indicates that some respiration-derived TDIC must originate from methane oxidation. d13C value of added TDIC is close to that of methane at these depths, indicating that methane oxidation is the dominant process by which TDIC has been produced.

here is clear; the shift to much lighter d13CTDIC can only be explained by increased TDIC production from respiration of contaminant carbon plus methane oxidation, with methane oxidation more dominant below 42 mbgl. At shallower depths in MLS60 there is a more subtle shift to increased TDIC production from respiration. At MLS59 the effect of the PAT system is more muted and not unidirectional. There is a slight shift towards greater TDIC production via respiration in most of the plume, whereas methanogenic TDIC production is more dominant at depth. 4.5. Implications and applications This study has shown that conventional PAT can be used to enhance the natural attenuation of organic contaminants in groundwater, by targeted intervention to improve conditions for in situ biodegradation. For the plume considered, biodegradation of the phenolic compounds was stimulated by reducing the contaminant concentrations below levels which inhibit biodegradation and increasing the flux of dissolved electron acceptors in background groundwater into the plume. This field-scale observation confirms the experimental results of Harrison et al. (2001), who showed that biodegradation rates of phenols in aerobic and anaerobic microcosms containing innocula from the aquifer increased with dilution of the contaminated groundwater. In principle, this engineered intervention can be applied to enhance the in situ restoration of biodegradable organic contaminants in many aquifers (US EPA, 1996) and therein has considerable potential. However, most plume management strategies which integrate pumping with in situ bioremediation focus on the addition of amendments (e.g. oxidants, nutrients or microorganisms) to increase biodegradation. This requires adequate mixing between contaminants and amendments in the plume or delivery of amendments to zones of interest to be effective. In this respect increasing the advection of background groundwater into a plume is key to using a PAT system to enhance the in situ biodegradation of oxidisable organic contaminants. Firstly, it introduces dissolved electron acceptors into the plume at a faster rate to meet the demand from the organic contaminants than under natural conditions. As these oxidants are typically more energetically favourable for biodegradation and bioavailable than mineral oxidants in aquifers (Wiedemeier et al., 1999), in situ biodegradation rates are likely to increase. This was shown by Harrison et al. (2001), who documented higher first-order biodegradation rates for the phenol compounds under aerobic rather than anaerobic conditions in laboratory microcosms. Secondly, transverse dispersion increases with groundwater velocity (Fetter, 1993), with the effect of mixing more dissolved oxidants into the plume from the background groundwater. Therefore, the PAT system should be designed to increase the advection and transverse dispersion of background groundwater into the plume to enhance in situ biodegradation. For the site studied, scenario

modelling indicated that this could also be achieved using alternate injection and abstraction of background groundwater into the aquifer via wells placed transverse within and outside the plume (Jones et al., 2002). This modelling analysis suggested that injection periods should be short to ensure adequate mixing with the plume and utilisation of oxidants in the injected groundwater. Under these conditions, predicted remediation timescales were considerable shorter than by natural attenuation alone. This engineered intervention is particularly useful in situations where the electron acceptor demand from the plume considerable exceeds the electron acceptor availability in the aquifer. Dispersion increases with increasing aquifer anisotropy and physical heterogeneity. Hence, a targeted PAT system is also likely to be a suitable basis to increase the natural attenuation of plumes in such cases. Modelling studies will be necessary to design and optimise the PAT system for this application, considering the abstraction well network (number, location, depth and screen interval of wells) and pumping regime (abstraction rates, well capture zones and management programme), as for the implementation of conventional PAT systems (US EPA, 1996; Suthersan, 1999). This analysis should also explore operational scenarios for the PAT system implementation, which maximise contaminant attenuation and remediation performance, considering relevant attenuation processes and parameters which describe these (Schaerlaekens et al., 1999; Park et al., 2007). 5. Conclusion Time-series studies of changes in the hydrochemistry and stable isotope composition of selected chemical species in groundwater demonstrate the potential of engineered restoration using pump and treatment (PAT) to enhance the biodegradation of organic contaminants in this aquifer. The PAT system reduced the organic contaminant concentrations in the plume below levels which had been shown in previous studies to suppress microbial respiration and limit contaminant biodegradation. At the same time, the targeted pumping regime enhanced the mixing between the plume and uncontaminated groundwater in the aquifer, increasing the supply of dissolved electron acceptors for anaerobic respiration in the plume. In turn, these combined effects (reduction in the dissolved organic contaminant load and increased influx of dissolved oxidants) stimulated the activity of the indigenous aquifer microorganisms and sulphate reducing bacteria in particular. This increased the relative importance of sulphate reduction at locations in the plume where this biodegradation process was previously suppressed. The enhancement of microbial respiration was marked by the increased production of dissolved inorganic carbon in the plume from organic contaminant and acetate biodegradation, and shifts in the stable isotope composition of residual sulphate, which confirmed the increased contribution of these

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microorganisms to contaminant turnover. Other respiration processes, such as anaerobic methane oxidation and denitrification, were also stimulated in situ, so that overall the potential for NA of the organic contaminants in the plume was increased during the operation of the PAT system. The cessation of pumping and recovery of dissolved organic contaminant concentrations towards pre-pumping values demonstrated the sensitivity of the in situ aquifer microorganisms to the hydrochemical conditions in the plume. This phase was marked by a shift in the balance between fermentation and respiration processes involved in contaminant biodegradation. The spatial distribution and relative contribution of respiration processes to biodegradation changed towards conditions resembling those before implementation of the PAT system, with a subsequent decrease in the importance of sulphate reduction. Fermentation processes (e.g. acetate production and acetoclastic methanogenesis) appear to be more robust and less affected by these changing conditions. This behaviour confirms that the indigenous aquifer microorganisms respond relatively quickly to perturbations in the plume organic chemistry induced by the PAT system. The activity and distribution of respiratory microorganisms may therefore be selectively stimulated by the PAT system, offering the potential to enhance overall contaminant biodegradation in the plume by appropriate intervention. Studying the shift in microbial community structure and succession during the operation of the PAT system will provide valuable insight on the microorganisms which contribute to these biodegradation processes and their metabolic function under the different conditions imposed by this engineered remediation. This study opens the possibility to integrate engineered restoration, such as groundwater pump and treatment, with in situ bioremediation or NA in a novel ‘‘dual-technology’’ management strategy for complex plumes. By targeting the pumping regime to focus on specific zones or conditions in plumes which limit in situ biodegradation processes, the performance of bioremediation or NA may be enhanced and the remediation timeframe shortened. This management concept could feasibly reduce the duration and associated cost of engineered restoration alone and clearly has potential application to many other contaminated aquifer settings than the context considered in the present study. Acknowledgements The authors gratefully acknowledge the support and assistance of the site owner in the completion of this research. This research was completed while KB was in receipt of an EPSRC PhD+ Fellowship and while LC was in receipt of a Marie Curie Early Stage Researcher Fellowship within the ADVOCATE Marie Curie Initial Training Network, funded by the European Commission (grant 265063). Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.apgeochem.2014. 06.023. References Ahad, J.M.E., Lollar, B.S., Edwards, E.A., Slater, G.F., Sleep, B.E., 2000. Carbon isotope fractionation during anaerobic biodegradation of toluene: implications for intrinsic bioremediation. Environ. Sci. Technol. 34, 892–896. Al-Khalid, T., El-Naas, M.H., 2012. Aerobic biodegradation of phenols: a comprehensive review. Crit. Rev. Environ. Sci. Technol. 42, 1631–1690. Allen, D.J., Brewerton, L.J., Coleby, L.M., Gibbs, B.R., Lewis, M.A., MacDonald, A.M., Wagstaff, S.J., Williams, A.T., 1997. The physical properties of major aquifers in England and Wales. British Geological Survey Technical Report WD/97/34. 312pp. Environment Agency, R&D Publication 8. Baker, K.M., Bottrell, S.H., Thornton, S.F., Peel, K.E., Spence, M.J., 2012. Effect of contaminant concentration on in situ bacterial sulfate reduction and

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