Identification of a reactive degradation zone at a landfill leachate plume fringe using high resolution sampling and incubation techniques

Journal of Contaminant Hydrology 85 (2006) 179 – 194 www.elsevier.com/locate/jconhyd

Identification of a reactive degradation zone at a landfill leachate plume fringe using high resolution sampling and incubation techniques Nina Tuxen, Hans-Jørgen Albrechtsen, Poul L. Bjerg Technical University of Denmark, Institute of Environment and Resources, Bygningstorvet 115, 2800 Lyngby, Denmark Received 5 July 2005; received in revised form 10 January 2006; accepted 19 January 2006 Available online 9 March 2006

Abstract Vertical small-scale variation in phenoxy acid herbicide degradation across a landfill leachate plume fringe was studied using laboratory degradation experiments. Sediment cores (subdivided into 5 cm segments) were collected in the aquifer and the sediment and porewater were used for microcosm experiments (50 experiments) and for determination of solid organic carbon, solid–water partitioning coefficients, specific phenoxy acid degraders and porewater chemistry. Results from a multi-level sampler installed next to the cores provided information on the plume position and oxygen concentration in the groundwater. Oxygen concentration was controlled individually in each microcosm to mimic the conditions at their corresponding depths. A highly increased degradation potential existed at the narrow plume fringe (37.7 to 38.6 masl), governed by the presence of phenoxy acids and oxygen. This resulted in the proliferation of a microbial population of specific phenoxy acid degraders, which further enhanced the degradation potential for phenoxy acids at the fringe. The results illustrate the importance of fringe degradation processes in contaminant plumes. Furthermore, they highlight the relevance of using high-resolution sampling techniques as well as controlled microcosm experiments in the assessment of the natural attenuation capacity of contaminant plumes in groundwater. D 2006 Elsevier B.V. All rights reserved. Keywords: Landfill leachate plume; Microbial degradation; Phenoxy acid; Plume fringe; High resolution

E-mail address: [email protected] (N. Tuxen). 0169-7722/$ - see front matter D 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.jconhyd.2006.01.004

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1. Introduction Old landfills without liners or leachate collecting systems constitute a threat to the downgradient groundwater quality due to infiltration of leachate into the subsurface. Often, high concentrations of organic carbon such as volatile fatty acids, humic like compounds and fulvic acids leach into the groundwater and form a leachate plume (Christensen et al., 2001). However, the greatest contamination threat is often posed by ammonium (Christensen et al., 2000a,b) or the many different xenobiotic compounds (e.g. BTEX compounds, phenolic compounds, chlorinated aliphatic compounds and a variety of pesticides) ¨ man and Hynning, 1998; Paxeus, 2000; Kjeldsen et al., found in leachate at landfill sites (O 2002; Baun et al., 2004). In a recent survey, pesticides were found in 9 out of 10landfill leachates, with phenoxy acids as the dominating compounds (Baun et al., 2004). The phenoxy acids and other xenobiotics only constitute a small fraction of the organic carbon in the landfill leachate (b 0.05–0.9%, Christensen et al., 2001). Still, from an environmental point of view the concentrations are high, with typical phenoxy acid concentrations of 10–250 Ag/L (Gintautas et al., 1992; Zipper et al., 1998; Baun et al., 2003); in one case up to several thousands of Ag/L (Williams et al., 2003)—far above the European Union drinking water standard of 0.1Ag/L. One of the major controlling parameters for the degradation of the phenoxy acids is the redox conditions. Oxygen influences the natural attenuation of these compounds particularly strongly. The phenoxy acids are mainly recalcitrant under anaerobic conditions, whereas aerobic degradation has been reported numerous times (reviewed by Reitzel et al., 2004), and a positive correlation between oxygen concentration and enhanced degradation (rates, mineralization, shorter lag phases) has also been established (Tuxen et al., 2005). The small concentrations of phenoxy acids in the leachate are not the controlling factor for the formation of anaerobic conditions found in the downgradient plumes. On the contrary, the formation of such conditions is mainly controlled by the degradation and subsequent consumption of electron acceptors by organic carbon leaching into the groundwater in high concentrations (Christensen et al., 2001). Accordingly, the fate of the phenoxy acids, which is highly dependent on the redox conditions, will be indirectly controlled by the fate of the organic carbon. It is well known that, in contaminant plumes, redox environments can vary greatly due to contaminant load, groundwater chemistry, geochemistry and microbiology inside the plume and along the flow path (Christensen et al., 2000a,b; Van Breukelen et al., 2003). Redox gradients (from highly reduced zones to oxidized zones) from the source towards the front of the plumes have been supported by detailed investigation of the terminal electron acceptor processes (Ludvigsen et al., 1999; Bekins et al., 2001). On the other hand the narrow plume fringes perpendicular to the flow direction which might be of special importance for the natural attenuation of oxidizable contaminants are less studied. Steep vertical concentration gradients for contaminants and redox parameters have been found in a few field studies of plume fringes, where contaminants mix with electron acceptors by dispersion and diffusion processes (Lerner et al., 2000; Thornton et al., 2001a; Van Breukelen and Griffioen, 2004). Furthermore, Pickup et al. (2001) found an increased phenol biodegradation potential at the plume fringe compared with either the uncontaminated aquifer or the plume core. From these observations, it could be hypothesized that the fringe zone hosts specific microbial populations able to degrade the contaminants in question and thereby a significant degradation potential. Microbial populations able to degrade specific organic chemicals can be quantified by e.g. the most probable number (MPN) culturing technique (e.g. Tora¨ng et al., 2003) or by detection of specific genes using PCR (poly chain reaction) amplification (e.g. Pickup et al., 2001 or de

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Lipthay et al., 2003). Small-scale variations in degradation potential can be efficiently investigated by laboratory microcosm experiments (Nielsen and Christensen, 1994). Experiments describing relevant features at the fringe should take the field scale gradients into account. This puts a burden on sampling techniques, resolution and number of samples. In order to obtain a realistic assessment of the degradation potential in contaminated plumes, it is important to collect aquifer material for the laboratory tests by use of high-resolution sampling techniques. The degradability could depend highly on electron acceptors and contaminant concentrations; these should therefore correspond to field conditions. Of special interest for aerobic degradable compounds such as phenoxy acids are the oxygen concentrations which will be a special challenge to control. The aims of this study were to determine across a landfill leachate plume fringe contaminated by phenoxy acid herbicides: (1) which factors (redox conditions, specific microorganisms and phenoxy acid concentrations) controlled the phenoxy acid degradation; (2) whether a zone existed at the plume fringe with an increased degradation potential; and (3) if the fringe zone harbored an increased number of specific phenoxy acid degraders. Parallel to the present study of the degradation potential, a study of the aquifer chemistry and natural attenuation capacity in the landfill leachate affected aquifer was performed on the basis of comprehensive field studies. 2. Materials and methods 2.1. Field site This study was performed with sediment and groundwater from a landfill leachate plume in Sjoelund, Denmark. The site history, hydrogeology and groundwater quality have been studied in a previous investigation (Tuxen et al., 2003). Sjoelund landfill was in use from 1965 to 1975 and received household waste, demolition waste and almost certainly some chemical waste. The landfill covers an area of 6300 m2 and has an average height of 5 m. Neither liners nor leachate collection systems exists. The landfill is located on a glacial outwash plain in a farmland area (Fig. 1). A 15–18 m thick unsaturated zone consisting of different glacial sediment intersects the landfill from a 3–5 m thick saturated aquifer. This aquifer is composed of fine-grained sand with silty and clayey lenses and is confined at the bottom by a 4–5 m thick aquitard (diluvial clay) separating the upper aquifer from a lower aquifer. The hydraulic gradient is 0.015 to 0.035 and the groundwater flow velocity is 30 to 80 m/year. The pristine part of the aquifer is aerobic (O2, 1–8 mg/L) with high nitrate concentrations (up to 35mg/l), and a non-volatile organic carbon content (NVOC) of 1–3mg/L. There is no significant presence of reduced species such as Fe2+, Mn2+, NH4+, HS and CH4, and the pH is neutral (7.0–7.5). Leaching of NVOC in mg/L concentrations from the landfill has resulted in the development of an anaerobic plume with slightly elevated Fe2+ and Mn2+ concentrations and depletion of nitrate and oxygen. Different phenoxy acid herbicides including MCPP and dichlorprop present in high concentrations (up to 65 Ag/L) were the compounds of major concern. 2.2. Collection and handling of groundwater and sediment Sediment for the study was collected as undisturbed cores across the landfill leachate plume fringe. In order to fulfill the need for sufficient amounts of sediment from each depth and a high

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Fig. 1. Sjoelund landfill with locations of the multi-level sampler and the two sediment cores.

vertical resolution at the same time, it was necessary to collect two sediment cores next to each other (1 meter apart, Fig. 1). Core X was 2.7m long (from 18.3 to 21 m below ground level, corresponding to 36.7–39.4 m above sea level, masl or 0.5 to 3.2 m below the ground water table) and core Y was 2.3 m long (from 18.8 to 21.1 m below ground level, corresponding to 36.6–38.9 masl or 1.0 to 3.3 m below the ground water table). A hollow stem auger drilled down to the desired top level of the cores. After release of the drill tip, the stainless steel sediment sampler (with an inner diameter of 57 mm) was pushed down in the undisturbed aquifer by an electrical hammer, then a further 5–10 cm down in the clay layer underlying the sandy aquifer. The clay acted as a stopper preventing the sediment and groundwater from being lost when the cores were withdrawn. The cores were sealed in the ends with aluminum foil and plastic stoppers to maintain the redox conditions and stored at 108C until used less than 4days later. To obtain uncontaminated material, 10 cm of each end of the cores was removed. In the laboratory the stainless steel cores were cut in 5cm segments by use of a tube saw (Georg Fischer). The sections were immediately sealed with aluminum foil and transferred to an anaerobic glove box (Coy Laboratory Products, Inc.). The sediment from each core segment was aseptically homogenized after removal of stones larger than 4 mm, then divided into smaller parts for the microcosm experiments, porewater extraction, and sediment analyses. Sediment samples from core X were used in microcosms, porewater extraction, determination of MPN, solid organic matter (TOC), and grain size distribution. Sediment samples from core Y were used for porewater extraction, determination of grain size distribution and distribution coefficients (K d-

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values) (Table 1). The sediment for microcosms, porewater extraction, and specific phenoxy acid degraders determination was used immediately. The sediment for TOC determination was stored at 10 8C, the sediment for determination of the grain size distribution was dried at 105 8C for 24 h, and the sediment for sorption determinations was frozen. Fifteen milliliters of milliQ water was added to the homogenized sediment from each core piece (50–100 g), then shaken for 24h followed by a centrifugation of the slurries at 2000 rpm for 17 min in order to extract porewater. The supernatant was filtered with an Advantec MFS-25 PTFE 0.2Am filter and frozen immediately until it was analyzed for phenoxy acids and chloride concentrations. The water content in each sample was determined by comparing the weight before the addition of the 15 mL milliQ water and after the centrifugation and a drying. The original porewater concentrations were then calculated by correcting for the dilution caused by the milliQ water addition. A multi-level-sampler (MLS B1) was installed in the aquifer 50 m downgradient of the landfill border (Fig. 1) with 30 sampling points (10 cm screens) separated by 12.5 or 25cm covering the entire aquifer thickness. Groundwater was sampled in a double tube system by nitrogen pressure because of the thick unsaturated zone (N 15meter), which ruled out suction pumps. From each screen, a total of 150mL of groundwater was collected. Approximately half this amount was divided into different vials and preserved for subsequent analyses (among others chloride, NVOC and phenoxy acids, Table 1). The remaining 75mL was in the field flushed through an airtight 20mL chamber containing devices for oxygen and electrical conductivity measurements (Table 1). The last 20mL of the 75mL was contained in the chamber and stirred until all measured values were stable. 2.3. Microcosm degradation experiments High-resolution (5 cm) microcosm experiments were performed in order to study the degradation potential in detail across the landfill leachate plume. In total 50 incubations were performed in 50mL sterilized infusion glass bottles, each containing aquifer material from the 5 cm sediment samples from core X (Fig. 2, Table 1). In an anaerobic glove box (Coy Laboratory Products, Inc.), 40 g wet aquifer material (sediment and porewater) and 3.5mL milliQ water containing 45 Ag/L 14C-MCPP (Izotop, radiochemical purity N95%) were added to each bottle Table 1 Overview of analysis done on the ground/pore-water and sediment in the current study

Ground/pore-water Electrical conductivity Oxygen NVOC Chloride Phenoxy acids Sediment Grain size Water content TOC MPN Kd Microcosm experiments

MLS

Core X

Core Y

x x x x x

x x

x x

x x x x

x x

x x

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Fig. 2. Microcosms for determination of degradation potential. See detailed description under Section 2.3.

increasing the original MCPP concentration of approximately 15 Ag/L and the volume ratio between the water phase and the sediment phase by approximately 50%. To two of the bottles, HgCl2 was added (equivalent to a concentration of 100mg/L in the bottles) together with the MCPP solution to serve as abiotic controls. After the bottles were taken out of the anaerobic box, they were flushed with an 80/20 (volume-%) mixture of N2 and CO2 in order to remove the H2 traces from the anaerobic box. They were then closed with butyl rubber stoppers. Oxygen concentrations were controlled during the experiment in each bottle individually by adding pure oxygen. This was done to obtain concentrations similar to those observed in the multi-level sampler at the corresponding depths. Finally, 1 mL 2.5 M KOH was added to the glass vial inside the bottles by using a syringe pushed through the stoppers. This was in order to trap the evolved 14 CO2. The microcosms were incubated for 246 days at 10 8C in darkness to simulate aquifer conditions. At the setup and at each sampling event (17 times during the experiment), the oxygen concentrations were monitored and more oxygen was added if the desired concentrations were not reached. The novel non-invasive oxygen method that was used implied that oxygen could be measured without affecting the system and in principle, an unlimited number of times. The base in the traps was collected and replaced using sterile syringes at each sampling event. After 246 days, 3 mL 5M HNO3 was added to the microcosms in order to drive out any 14CO2 present in the water phase. Pressure built up as a result of dissolution of calcite in the sediment.

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Because of this, an external base trap containing 2mL 2.5 M KOH was connected to the bottles to supplement the capture of 14CO2 in the internal base traps. At the end of the experiment the sediment from each microcosm was dried and grinded. In order to establish a mass balance, triplicate samples of 300mg were collected to determine the remaining 14C in the sediment and the water phase. 2.4. MPN determinations The number of specific phenoxy acid degraders was enumerated by a most probable number (MPN) method relying on phenoxy acid degradation. Two milliliters of a minimal medium (ISO 7827) containing 29Ag/L 14C-MCPP was added to each of the 24 wells in a 6  4 multi-dish (Nunc), except in the upper row of the wells where a 10 times stronger media were added. From each of the 50 sediment samples (core X, Table 1), 20g was diluted in 40 mL 0.01 M P-buffer (containing NaH2PO4d H2O and Na2HPO4d 2H2O, pH 7.4), vortexed for 1 min and diluted in six 10-fold serial dilutions (100–10 5). The wells in the multi-dishes (one multi-dish per sediment sample) were then supplied with 4replicates of each sediment dilution—2 mL aliquot of the 100 dilution in the upper row and 200 AL aliquots of all dilutions in the remaining wells. The multidishes were closed with lids and incubated aerobically at 10 8C. After 7 months of incubation the multi-dishes were opened and 100AL concentrated HCl was added to each well to remove 14 CO2. The residual 14C-MCPP was quantified using liquid scintillation counting and MPN was scored positive if more than 25% of the added MCPP was mineralized. 2.5. Analytical procedures Oxygen concentrations were analyzed by the use of a fiber-optic oxygen technique developed by PreSens Precision Sensing GmbH, where oxygen concentrations are proportional to the luminescence decay time of a sensor foil. Oxygen concentrations in the water from the multilevel sampler were measured using a polymer optical fiber coated with an oxygen-sensitive foil covered by a steel tube placed inside the groundwater chamber. Oxygen in the microcosm experiments was measured from the outside of the microcosm by a fiber optical oxygen meter and using oxygen-sensitive luminescent sensor foil mounted inside the microcosms. The detection limit was 0.15 mg/L. Electrical conductivity was measured using a WTW LF 330 Conductivity Meter. The phenoxy acids (4-chloro-2-methylpropanoic acid (MCPP), 2,4-dichlorophenoxypropanoic acid (dichlorprop), 2-chlorophenoxypropanoic acid (2-CPP) and 4-chlorophenoxypropanoic acid (4CPP)) were analyzed on a Hewlett Packard Series 1100 HPLC system at wavelength 205nm with a detection limit of about 2Ag/L (Broholm et al., 2001). Chloride was measured on a Dionex ion chromatograph DX120 and NVOC was determined on an O.I. Model NVOC-analyzer. The 14C in the dried sediment, after the closure of the degradation experiment, was combusted for 15s in a Packard Model 307oxidizer with an excess of O2. The evolved 14CO2 was trapped in 8 mL Carbosorb E+ (Packard Bioscience). Ten milliliters Permaflour E+ scintillation cocktail was added to this fraction (Packard Bioscience) and the amount of 14 CO2 was quantified by a WinSpectralTM, 1414 Liquid Scintillation Counter. Similarly, 4 mL HiSafe 3 (Wallac) scintillation cocktail was added to the 14CO2 collected in the base traps and quantified by a WinSpectralTM, 1414 Liquid Scintillation Counter. TOC in the sediment (core X, Table 1) was determined in a LECOR-oven after removal of inorganic carbon by treatment with 6% H2SO3. The detection limit was 0.01 mg C/g dw. Grain

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size was analyzed using 6 different sieves covering the range of 0.063 mm to 2 mm. The fraction below 0.063 mm was differentiated by the use of X-ray sedimentation on a Micromeritics SediGraph 5100. K d-values were determined using a batch equilibrium procedure and calculated from the mass balance assuming a linear isotherm according to Clausen et al. (2001). All sediment samples (core Y, Table 1) were investigated in triplicate at 10 8C in darkness in sterilized glass tubes with Teflon caps. Five grams of freeze-dried sediment was equilibrated with 4 mL sterilized groundwater from the aquifer for 24 h. After which, another 1mL sterilized groundwater containing 50Ag/L 14C-MCPP was added, resulting in a concentration of 10 Ag/L. After 96 h of equilibration the glass tubes were centrifuged at 1500rpm for 10min and 1mL supernatant was collected. The amount of 14C-MCPP was determined by liquid scintillation counting after the addition of 10 mL OptiPhase HiSafe 3 scintillation cocktail. 3. Results and discussion 3.1. Leachate plume position Landfill leachate contains elevated electrical conductivity in comparison with pristine groundwater. Since the uncertainty in background levels is normally low, and breakthrough of the electrical conductivity in most cases is not retarded compared to the groundwater, the electrical conductivity is considered to be an ideal indicator of the position of an overall leachate plume (Christensen et al., 1992). Chloride can also be used but variability in background concentrations can make interpretation difficult (Bjerg and Christensen, 1992; Bjerg et al., 1995). In the upper screens of MLS B1 down to 38.5masl, electrical conductivities were similar to background values in the aquifer (Tuxen et al., 2003). However, below 38.5masl the values increased until the highest values of 1200–1250 AS/cm were reached, 36.5–37.7 masl (Fig. 3A). These results revealed that the core of the plume at the MLS B1 position was located between 36.5 and 37.7masl, surrounded by a fringe up to 38.5masl. A comparison of the chloride (Fig. 3B) and MCPP concentrations measured in MLS B1 and the concentrations in the porewater from the two cores (data not shown) showed that the leachate plume was positioned at exactly the same depth at the three sampling positions regardless of the undulating surface of the fine-

Fig. 3. Vertical groundwater chemistry profiles for multi-level sampler B1 at October 2003. (A) The electrical conductivity and the oxygen concentration. (B) The concentration of chloride and NVOC. The groundwater table was located at 39.9masl.

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grained sediment. Thus, the use of results from multiple sources (two cores and one MLS) as a representation of the fringe seems justified. Oxygen concentrations in MLS B1 (Fig. 3A) revealed an anaerobic plume core. This is in contrast to the background conditions above the plume where the oxygen concentrations were high (up to 6 mg/L) due to a continuous infiltration of oxygenated recharge. Between these extremes, the vertical concentration gradient was steep. The explanation for this oxygen distribution is that NVOC from the landfill has depleted the oxygen in the core of the plume in the otherwise aerobic aquifer (Fig. 3B, Tuxen et al., 2003). 3.2. Characterization of the sediment The upper parts of the two sediment cores consisted of sand with a few percent clay and silt and two narrow zones with higher gravel contents (core X is shown in Fig. 4). Below 37.3m above sea level (masl) both sediment cores had a significant higher content of fine grained material, with up to 24% clay and 66% silt. The similar grain size distribution in the two cores confirms that results from the two cores can be interpreted as originating from one core with respect to geology. The organic carbon content (TOC) of the sediment ranged from 0.1 to 1.2mg C/g dw and did not show a systematic vertical variation related to the plume position (Fig. 4). These values are in the same range as TOC concentrations found in other Danish aquifers (Christensen et al., 1996; Clausen et al., 2004). In general, the fraction of fine grained material correlated with TOC (R 2 = 0.9 using a logarithmic correlation) excluding the three samples with a gravel content above 5%. The higher organic carbon content (0.6–1.8 mg C/g dw) in these three samples is probably a result of the removal of stones during preparation of the samples, with the consequence that the relative weight of the organic carbon was greater compared to samples where a complete homogenization was possible. Such a correlation between TOC content and the smallest particle-size fraction has been shown in other studies (Madsen et al., 2000; Clausen et al., 2004) and is thought to result from the high sorption affinity of TOC to clay minerals (Stevenson, 1994) and/or from the apparent resistance to biodegradation of humus bound to clay minerals (Sposito, 1989). K d-values for MCPP were in the range of 0.02–0.11 L/kg (a single value of 0.21 L/kg) (data not shown). The values were in agreement with those found in similar aquifers (Tuxen et al.,

Fig. 4. Grain size distribution and organic carbon content in the sediment core X.

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2000) where the pH values (5.3 to 7.1) resulted in an almost full dissociation of the phenoxy acids. There was no correlation between the K d-values and the depth, the TOC content and the grain size distribution. A correlation with the TOC content was not expected since dissociated compounds are extremely water soluble and thus not susceptible to hydrophobic partitioning with the organic fraction of the sediment (Schwarzenbach et al., 1993). On the other hand, a correlation with the fine grained fraction was expected, since it has been shown that phenoxy acids bind to clay minerals via electrostatic interactions (Clausen et al., 2001). The measured K dvalues correspond to low retardation factors, R, of 1.1–1.6 (using porosities and bulk densities estimated from the water content and a presumed particle density of 2.65 kg/L) and hence sorption seems not to be a controlling factor of MCPP fate in the aquifer. 3.3. Experimental conditions Analyses of phenoxy acids in the porewater from core X revealed elevated concentrations in the part of the aquifer where the landfill leachate plume was located (Fig. 5A). This corresponds with the electrical conductivity and the chloride measured in the MLS as well as the porewater from the two cores. The dominant phenoxy acid in the core was MCPP with concentrations up to 220 Ag/L and only relatively small concentrations (b15 Ag/L) of the other phenoxy acids dichlorprop, 2-CPP and 4-CPP (data not shown). The porewater was extracted from subsamples of the same batch used for the microcosm experiments and thus reflected the initial MCPP concentration in each of the bottles. The only difference was that 10Ag/L 14C-MCPP was added to the bottles during the setup. Calculations using the measured K d-values for each depth and the specific water and sediment content in the microcosms at the corresponding depth resulted in MCPP being present for more than 85% in the water phase. Accordingly, the MCPP would, independently of the desorption rate, be available for the microorganisms as long as the mineralization was less than 85%. To provide as realistic conditions as possible for the biodegradation potential assessments, the conditions in the aquifer were simulated in the microcosms by adjusting and controlling the oxygen concentrations of each bottle throughout the experimental period. This was achieved by

Fig. 5. (A) MCPP concentrations in the porewater from the sediment core X (used for the microcosm experiments) and the average oxygen concentrations in each microcosm following an initial stabilizing period of 8days (Fig. 6A) and until closure of the experiments after 248days. (B) The initial number of specific MCPP degraders in the sediment evolved CO2 as well as the total amount of MCPP
degraded during the experimental period calculated as Mdegraded ¼ ð1Y Þ Cporewater dVporewater þ Cadded dVadded , assuming a yield coefficient, Y, of 0.3 g C in biomass/g C degraded as found in Tuxen et al. (2002).

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Fig. 6. (A) Oxygen concentrations in 4 of the 50 microcosm experiments during the experimental period of 248 days. The arrows indicate the aimed oxygen concentration corresponding to the field measurements. (B) Evolved 14CO2 as % of the amount of 14C-MCPP initially added in 4 of the microcosm experiments during the experimental period of 248days.

using novel non-invasive oxygen analysis in the experiments. The aim of achieving the same oxygen concentration in each microcosm as was found in the corresponding depth in the MLS was successful (compare Figs. 3A and 5A). However, an initial period of 8 days with frequent oxygen-amendments was necessary in order to reach the desired levels (Fig. 6A), since all samples were treated anaerobically until this stage to avoid any oxygen in the original anaerobic samples. After the initial period of 8 days, a fairly constant oxygen concentration which paralleled the field conditions was maintained by means of occasional rectifications (Fig. 6A). 3.4. Degradation potential for phenoxy acids The phenoxy acid degradation potential was assessed by monitoring the 14CO2 evolved in each microcosm during the experimental period (Fig. 6B). In approximately half of the microcosms 14CO2 evolved above the radioactive impurity content and thus confirmed that mineralization of the added 14C-MCPP occurred. No mineralization occurred in the abiotic controls. The remaining 14C in the water and sediment phase constituted between 70 and 117% of the added 14C resulting in a total average 14C mass balance of 91% (standard deviation 13%) with no systematic vertical variation (Fig. 7). Large differences in mineralization were observed between the microcosms covering the range from no mineralization to mineralization of 30% of the added 14C-MCPP (4examples are shown on Fig. 6B). Mineralization was only observed in microcosms with sediment from the

Fig. 7. Final

14

C mass balance for microcosm experiments for all experiments performed with sediment from core X.

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narrow zone between 37.6 and 39.0 masl and the highest values were found in an even smaller zone of less than 1meter between 37.7 and 38.6masl. In some microcosms mineralization started immediately whereas in others, a lag period of up to 80–100 days (dependent on the definition of the lag period) preceded mineralization. An apparent positive relationship between the evolved 14CO2 (and thus degradation potential) and oxygen concentration was observed in a large number of the microcosms (illustrated by microcosms A, B and C in Fig. 6A and B), as confirmed by the literature (Tuxen et al., 2005; Vink and Van der Zee, 1997). Similarly, the length of the lag periods decreased with increasing oxygen concentrations, which has also been observed by Tuxen et al. (2005). However, in the microcosms with the highest oxygen concentrations (N 4–5 mg/L,) the degradation potential decreased and the length of the lag periods increased (illustrated by microcosm D in Fig. 6A and B). Hence, oxygen was not the only controlling factor for the degradation potential. The total amount of MCPP degraded during the experimental period was calculated assuming that the evolved 14CO2 (above the radioactive impurity) from the added 14C-MCPP reflected the degradation of the original MCPP in the porewater (Fig. 5B). Presenting the degradation data in this form highlighted the observations from the individual mineralization curves: large variations in the degradation potential over small vertical distances were observed. In the narrow zone between 37.7 and 38.6 masl the degradation potential was much higher than in the zones below and above. 3.5. Specific degraders The number of specific aerobic MCPP degraders in the sediment also varied with depth (Fig. 5B) with very few degraders (0.5–36 cells/g dw) in the upper part of the core (38.6–39.3 masl). Similarly, there were also few degraders (2–600 cells/g dw) in the lower part of the core (36.8– 37.7 masl). In between, the number of MCPP degraders was elevated with 200–3600 cells/g dw. The highest numbers of specific MCPP degraders were of the same order of magnitude as the number of specific MCPP degraders determined in another aerobic MCPP exposed aquifer (de Lipthay et al., 2003; Tora¨ng et al., 2003). Additionally, these studies have shown that the specific MCPP degraders only constituted b 1x of the total bacterial population (determined using DAPI-staining). In the lower, anaerobic part of the aquifer, MPN numbers were slightly elevated in some samples, but no degradation took place in the microcosms with aquifer material from this zone. The explanation should probably be found in the incubation circumstances: the MPN multidishes were incubated aerobically, and in the lower part of the aquifer the MCPP concentrations were high in the porewater. As a result the microbial population adapted to degrade MCPP in some samples. Other studies have shown that originally anaerobic microbial populations can adapt to degrade phenoxy acids aerobically after the addition of oxygen (Tuxen et al., 2005). In contrast, the microcosms with aquifer material from the lower part of the aquifer contained no oxygen or very low oxygen concentrations which strongly limited the degradation. 3.6. Reactive degradation zone at the plume fringe Overall, a highly increased degradation potential expressed as the amount of MCPP degraded in the microcosm experiments was observed in a narrow zone between 37.7 and 38.6 masl. Furthermore, this zone harbored a significantly increased number of specific degraders (Fig. 5B). The increased degradation and decreasing lag periods correlated positively with oxygen

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concentrations and the number of specific MCPP degraders. Furthermore, this zone corresponded exactly to the location of the plume fringe determined by the electrical conductivity and chloride in the MLS, as well as the concentrations of MCPP in the porewater (Figs. 3 and 4B). Thus, optimal conditions for degradation of phenoxy acids have developed at the fringe, where MCPP and oxygen co-exist. As a result, populations of microorganisms capable of degrading phenoxy acids have proliferated, which in turn have further increased the degradation potential at the fringe. This is in agreement with previous studies by Tuxen et al. (2002), de Lipthay et al. (2003) and Tora¨ng et al. (2003) who demonstrated that degradation of phenoxy acids correlated closely to the number of specific degraders, which were also linked to pre-exposure to these compounds. Slight deviations from the abovementioned correlation between phenoxy acid pre-exposure, the degradation potential and the number of specific degraders were observed. In the lower, anaerobic part of the aquifer where no degradation was observed, slightly elevated MPN numbers were found in some samples, which, as mentioned previously, could be attributed to the differences in incubation circumstances. In the sediment from the upper part of the aquifer (above 38.6 masl) some degradation was observed despite the very low MPN numbers. These MPN numbers, however, were drawn from aquifer samples with only the original MCPP content in the porewater, which, in the upper part was below our detection limit. Re-analysis of the samples with the method described by Reitzel (2005), which had a detection limit below 0.02Ag/L confirmed this. In contrast, 0.16 Ag 14C-MCPP corresponding to 15Ag/L was added to the microcosm. This amount was partly degraded during the incubations after adaptation of the microbial population and corresponds with the observation of increased lag periods in samples above the fringe. Fringe degradation is not only important in the case of MCPP or other phenoxy acids in landfill leachate plumes. Benzene at the Vejen Landfill site, DK also behaved conservatively in the anaerobic part of the plume, and would most likely be degraded at the fringe between the anaerobic core and aerobic aquifer Baun et al. (2003). Van Breukelen and Griffioen (2004) demonstrated degradation of DOC at the fringe of the Banisveld landfill in the Netherlands. Plumes with BTEX compounds or MTBE also have the potential to be fringe controlled as documented as such at several sites, although degradation of BTEX in the core may be substantial as well (Cozzarelli et al., 1999; Gieg et al., 1999; Bekins et al., 2001; Reitzel, 2005). The well characterized Four Ashes site, UK with phenolic compounds is suggested to be fringe controlled (Thornton et al., 2001a,b). Ammonium plumes have also been speculated to show the same features (Maier and Grathwohl, 2005). Thus, the fringe as a highly reactive zone is an important topic. The proliferation of specific degraders and an enhanced degradation potential may be general features for many contaminant plumes. 4. Conclusions Steep vertical gradients in electrical conductivity and concentrations of chloride, oxygen and phenoxy acids were observed across a landfill leachate plume. Electrical conductivities, chloride and phenoxy acid concentrations were higher, while oxygen concentrations in the plume were lower than in the overlying uncontaminated groundwater. This defined a narrow fringe zone of less than 1 m thickness. At the fringe of the plume, there was a highly increased phenoxy acid degradation potential which could be attributed to the co-existence of phenoxy acids and oxygen. This resulted in the development of a microbial population with a specific phenoxy acid degrading capacity which further enhanced the degradation potential at the fringe.

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