Biological transformations of 1,2-dichloroethane in subsurface soils and groundwater

Journal of Contaminant Hydrology 34 Ž1998. 139–154

Biological transformations of 1,2-dichloroethane in subsurface soils and groundwater G.M. Klecka ˇ ) , C.L. Carpenter, S.J. Gonsior EnÕironmental Chemistry Research Laboratory, Health and EnÕironmental Research Laboratories, Dow Chemical, 1803 Building, Midland, MI 48674, USA Received 21 October 1997; accepted 29 May 1998

Abstract The ability of naturally occurring microorganisms to biodegrade 1,2-dichloroethane was examined in soilrwater microcosms prepared using aquifer material obtained from manufacturing sites in Louisiana and Texas with known histories of exposure to the compound, as well as in aquifer samples taken from a site in Oklahoma with no known history of 1,2-dichloroethane contamination. Biotransformation of 1,2-dichloroethane was noted under methanogenic or sulfate reducing conditions in all samples. Under anaerobic conditions, 1,2-dichloroethane was transformed to ethylene in a single step via reductive dihaloelimination. No other metabolites were detected in the reaction mixtures. Microbial adaptation appeared to be required for biotransformation of 1,2-dichloroethane. Lag periods ranging from 7 to 8 weeks preceded degradation in microcosms prepared with aquifer material from the Texas and Oklahoma sites. In contrast, no lag period was evident prior to biotransformation in microcosms prepared from the Louisiana manufacturing site, which is consistent with field evidence for natural biological attenuation in situ based on analysis of the groundwater chemistry. Aerobic biodegradation of 1,2-dichloroethane to carbon dioxide was also observed after 13 weeks in aquifer material from the Louisiana site, but was not evident in samples from the Texas or Oklahoma sites following 18 weeks of incubation. The ability of naturally occurring microorganisms to degrade 1,2-dichloroethane has bearing on assessments of the fate and lifetime of the compound in the environment, as well as having potential application in the remediation of contaminated groundwater. q 1998 Elsevier Science B.V. All rights reserved. Keywords: 1,2-Dichloroethane; Ethylene dichloride; Biotransformation; Reductive dehalogenation; Natural attenuation; Bioremediation; Groundwater

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0169-7722r98r$19.00 q 1998 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 7 7 2 2 Ž 9 8 . 0 0 0 9 6 - 5

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1. Introduction The widespread use of 1,2-dichloroethane as a chemical intermediate, solvent and lead scavenger in gasoline has stimulated interest in the behavior of the compound in the environment. Because of the relatively high water solubility Ž8524 mgrl; Horvath, 1982. and the potential for migration in soil ŽHoward, 1990., 1,2-dichloroethane has been detected in groundwater ŽPlumb, 1987.. Because groundwater often moves slowly, and residence times can range from years to centuries, the fate and lifetime of 1,2-dichloroethane in subsurface environments is of interest. 1,2-Dichloroethane has been shown to be susceptible to both abiotic and biological transformations ŽStucki et al., 1983; Janssen et al., 1985; Belay and Daniels, 1987; Egli et al., 1987; Stucki et al., 1987; Vandenbergh and Kunka, 1988; Barbash and Reinhard, 1989; Jeffers et al., 1989; Holliger et al., 1990a,b; van den Wijngaard et al., 1992.. Abiotic degradation has been well documented; vinyl chloride was detected upon alkaline hydrolysis of 1,2-dichloroethane, whereas reactions at neutral pH favor a hydrolytic substitution reaction, yielding ethylene glycol as the product ŽJeffers et al., 1989.. The half-life for the reaction at pH 7 and 258C has been reported to be 72 years ŽJeffers et al., 1989.. However, the abiotic reaction rate can be enhanced by the presence of certain anions frequently encountered in aqueous environments. For example, Barbash and Reinhard Ž1989. reported the half-life Ž258C. of the compound decreased to 37 years in the presence of 50 mM phosphate buffer, and to 6 years in reactions containing 50 mM phosphate buffer and 0.67 mM sodium sulfide. The aerobic biodegradation of 1,2-dichloroethane has been extensively studied ŽStucki et al., 1983; Janssen et al., 1985; Vandenbergh and Kunka, 1988; van den Wijngaard et al., 1992., and has been shown to result in complete mineralization of the compound to carbon dioxide, water, and inorganic chloride. Stucki et al. Ž1983. initially described the isolation of an unidentified bacterium, strain DE2, that was capable of utilizing 1,2-dichloroethane as a sole carbon source for growth. Janssen and coworkers ŽJanssen et al., 1985; van den Wijngaard et al., 1992; Janssen et al., 1995. have described the isolation and characterization of strains of Xanthobacter autotrophicus and Ancylobacter aquaticus with the ability to grow on 1,2-dichloroethane. The initial reaction is catalyzed by a haloalkane dehalogenase which converts 1,2-dichloroethane to 2-chloroethanol. Subsequent degradation involves a series of two sequential oxidations catalyzed by alcohol and aldehyde dehydrogenases to yield chloroacetate, which is finally converted to glycolate by the action of a haloacetate dehalogenase. Based on results of extensive biochemical as well as genetic analysis of the enzymes involved, the authors speculate that evolution of organisms with the ability to grow on 1,2-dichloroethane requires a number of steps ŽJanssen et al., 1995.. While two of the enzymes, namely the alcohol dehydrogenase and the chloroacetate dehalogenase appear to be common in nature, the haloalkane dehalogenase and chloroacetaldehyde dehydrogenase appear to be specifically adapted for the degradation of xenobiotic substrates. The biotransformation of 1,2-dichloroethane has also been reported under anaerobic conditions. Belay and Daniels Ž1987. and Egli et al. Ž1987. have described the biotransformation of 1,2-dichloroethane to ethene by pure cultures of sulfate reducing or methanogenic bacteria. In contrast, Holliger et al. Ž1990a. observed that cell suspensions

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of methanogenic bacteria reductively dechlorinated 1,2-dichloroethane via two different reaction mechanisms; a dihaloelimination reaction yielding ethene as well as two consecutive hydrogenolysis reactions yielding chloroethane and ethane. Stimulation of methanogenesis caused an increase in the amount of dechlorination products formed, whereas the opposite was found when methane formation was inhibited. Although the biotransformation of 1,2-dichloroethane is well documented, the majority of the work has involved the use of pure or enrichment cultures. Few data are available that can be used to predict the fate and lifetime of the compound in groundwater. Consequently, the present investigation was conducted to examine the biotransformation of 1,2-dichloroethane by naturally occurring microorganisms in aquifer materials. In addition to examining the pathways involved in degradation, the study was designed to evaluate factors that may influence reaction rates by examining biotransformation under a variety of environmental conditions. 2. Description of the study sites The ability of naturally occurring microorganisms to degrade 1,2-dichloroethane was examined in aquifer samples obtained from two different manufacturing sites with known histories of exposure to the compound. For comparison, biodegradation was also examined in aquifer samples taken at a site with no known history of 1,2-dichloroethane contamination. 2.1. Louisiana site Soil and groundwater samples from a 1,2-dichloroethane contaminated aquifer were collected at a manufacturing site in Plaquemine, LA. The soils which extend from the surface down to a depth of approximately 9 m have been characterized as a shallow permeable zone, with a water table at 0.3 to 0.6 m below grade. The upper 1 to 2 m consists primarily of coarse sand, which overlies a 6- to 9-m layer of silty clay. For the present investigation, samples of both the sand and silty clay were collected from locations which exhibited 1,2-dichloroethane concentrations in the range of 50 to 150 ppm. Prior work revealed correlations between organic carbon and inorganic chloride concentrations in groundwater samples obtained from the shallow permeable zone, suggesting that biodegradation might be occurring naturally in the aquifer ŽKlecka ˇ et al., 1994.. High inorganic chloride concentrations, ranging from 135 to 477 ppm, were detected in groundwater samples with high organic carbon concentrations Ž301 to 684 ppm.. In contrast, groundwater samples collected from an up gradient location indicated average background levels of organic carbon and inorganic chloride were 5 and 23 ppm, respectively. 2.2. Texas site Aquifer material with a history of exposure to 1,2-dichloroethane was also obtained from a manufacturing facility in Freeport, TX. The stratigraphy of the site is character-

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ized by a transmissive zone of fine sand 5 to 10 m thick, which is sandwiched between upper and lower clay layers that range in thickness from 5 to 12 m. Samples of the sand layer were collected from a region which exhibited 1,2-dichloroethane levels in the range of 50 to 100 ppm. 2.3. Oklahoma site Aquifer samples with no known history of exposure to 1,2-dichloroethane were collected from a shallow aquifer in Norman, OK. The study site has been previously described ŽDunlap et al., 1976; Beeman and Suflita, 1987; Klecka ˇ et al., 1990., and is located along the north bank of the South Canadian River and adjacent to the Norman municipal landfill. The aquifer at this location has been characterized as a highly permeable, alluvial aquifer composed of sand, silt, clay and gravel. The depth of the alluvium varies from 10 to 14 m and lies over a 100-m layer of dense clay and chert gravel. The water table averages from about 0.6 to 1.5 m below the soil surface ŽRobertson et al., 1974.. One of the sampling sites was located in an anaerobic portion of the aquifer that receives leachate from the municipal landfill. Suflita and coworkers ŽBeeman and Suflita, 1987; Adrian et al., 1994. have previously characterized this site as methanogenic based on the dominant microbial processes involved in carbon dissimilation. However, seasonal variations in sulfate concentrations in the groundwater are known to cause shifts between methanogenesis and sulfate reduction. In addition to the anaerobic site, shallow aquifer soil from a second location was obtained to evaluate the potential for biodegradation under aerobic conditions. The latter site was located to the north and out of the zone of influence of the landfill ŽKlecka ˇ et al., 1990..

3. Materials and methods 3.1. Test chemicals Radioactively labeled 1,2-dichloroethane Ž1,2-14 C; 6.8 mCirmmol. was purchased from DuPontrNEN Products, Boston, MA. The test chemical was shown to be 97% radiochemically pure by high pressure liquid chromatography as outlined below. Highpurity 1,2-dichloroethane Ž99.9%. was obtained from the Aldrich Chemical, Milwaukee, WI. Ethylene Ž99.5%. was purchased from Scott Specialty Gases, Troy, MI. Resazurin Žtechnical grade. was from the Sigma Chemical, St. Louis, MO. Chromatographic grade acetonitrile and isooctane were from the Fisher Scientific, Pittsburgh, PA. All other chemicals were reagent grade. 3.2. Aquifer materials Aquifer materials were collected from the sampling sites previously described. Shallow soils ŽLouisiana and Oklahoma sites. were obtained by digging to the top of the water table with a shovel or backhoe, and hand-filling sterile glass jars to capacity.

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Groundwater was then collected by digging a hole approximately 1 m deep, allowing the hole to fill, and then bailing water into sterile 1-gal Ž3.8 l. glass bottles. A hollow stem auger and split spoon sampler were used when collecting aquifer soils from deeper sampling locations ŽLouisiana and Texas sites.. Following sampling, the contents of the sampler were transferred to sterile glass jars. Groundwater from these locations was obtained from existing monitoring wells using a pump. All aquifer samples were chilled on ice and shipped to the laboratory by overnight express carrier. 3.3. Biodegradation experiments Biodegradation was examined under both anaerobic and aerobic conditions in microcosms constructed with combinations of subsurface soil and groundwater from the various sites. Anaerobic microcosms were prepared by aseptically transferring 40 g of soil Žwet weight. and 30 ml of filtered Ž0.45 mm. groundwater into sterile 70-ml serum bottles. The groundwater was purged with N2 gas prior to use, and the reaction mixtures were prepared in an anaerobic glove box containing an atmosphere of 70% N2r28% CO 2r2% H 2 . Resazurin Ž0.0002%. was added as the redox indicator. Aerobic biodegradation was examined in a similar manner, except that microcosms were prepared in a fume hood. To ensure maintenance of aerobic conditions, the microcosms were purged with O 2 gas for several minutes. Reaction mixtures were amended with 25 ml of a 1,4-dioxane stock solution containing 2.0 mg of w14 Cx1,2-dichloroethane. Because the aquifer material from several of the sampling sites was known to be contaminated, the initial concentration of 1,2-dichloroethane in the microcosms varied among the different experiments. All reaction mixtures were sealed with Teflon-faced butyl rubber septa and aluminum crimp seals, and incubated in an inverted position under static conditions at 258C in the dark. Killed controls were included in the study to monitor for nonbiological losses of the test chemical. Reaction mixtures were prepared as described above, and were adjusted to contain 2% formalin. Reaction mixtures were periodically sacrificed and analyzed for disappearance of w14 Cx1,2-dichloroethane and formation of products. Prior to analysis, the color of the redox indicator was noted and the samples were chilled on ice for 30 min. Reaction mixtures were extracted with 6 ml of isooctane for 30 min on a reciprocating shaker. A portion of the isooctane extract was analyzed by high pressure liquid chromatography and liquid scintillation counting as outlined below. The aqueous fraction was filtered through a 0.45-mm Acrodisc membrane filter ŽGelman Sciences, Ann Arbor, MI. and analyzed in an identical manner. Mineralization of radioactively labeled test substrates to w14 Cxcarbon dioxide was also monitored during the course of the study. w14 Cxcarbon dioxide was collected by passing nitrogen gas at a rate of 250 mlrminute through reaction mixtures which had been acidified by the addition of 1 ml of concentrated phosphoric acid. The effluent gas was collected in a series of two traps, each containing 10 ml of 1 N KOH solution. The trap solutions were combined and portions were analyzed by liquid scintillation counting. The production of w14 Cxcarbon dioxide was confirmed by adding 2.5 g of BaŽNO 3 . 2 to the trap solution, mixing for 30 min, and assaying the radioactivity in the solution after membrane filtration of the barium carbonate precipitate formed ŽRapkin, 1962..

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3.4. Analytical methods The distribution of w14 Cx-labeled 1,2-dichloroethane and any degradation products in the organic extract and the aqueous portions of the soil microcosms was determined by high pressure liquid chromatography. Samples were analyzed using a system composed of a Rheodyne model 7125 injector and Waters model 510 solvent delivery system, and an LDC SpectroMonitor III variable wavelength detector adjusted to 195 nm. Separations were achieved on a ZORBAX ODS reverse phase analytical column Ž4.6 mm = 25 cm; DuPont, Wilmington, DE. using a solvent system composed of acetonitrile–water Ž50:50.. The flow rate was 1 mlrmin at approximately 1000 psi. Effluent from the column was analyzed using a Berthold model 506A on-line radioactivity monitor. Total radioactivity in the organic and aqueous fractions of the reaction mixtures was determined by liquid scintillation counting. Samples were transferred to 10 ml of Aquasol scintillation cocktail ŽDuPontrNEN Products. and analyzed using a Beckman model 7800 liquid scintillation counter. Methane and ethylene were analyzed by direct injection of a portion of the headspace into a Hewlett Packard model 5710 gas chromatograph equipped with a flame ionization detector. The detector output was connected to a Hewlett Packard laboratory automation system. The analytical column Žglass; 1.83 m = 2.0 mm i.d.. was packed with Carbosieve S Ž100r120 mesh; Supelco, Bellefonte, PA.. Compounds were eluted from the column isothermally at 1008C with helium as the carrier gas at a flow rate of 20 mlrmin. Hydrocarbon concentrations were determined by comparison of the detector response to that of an external standard prepared in air. w14 Cx-labeled methane and ethylene were determined by bubbling the effluent gas from the flame ionization detector through 10 ml of 1 N KOH. Portions of the trap were transferred to 10 ml of Aquasol scintillation cocktail and analyzed as previously described. Soil and ground water samples were submitted for analysis to A and L Midwest Agricultural Labs, Omaha, NE. Soil samples were analyzed for texture and organic and inorganic content according to conventional methods ŽBlack et al., 1965.. Soil moisture was determined by gravimetric analysis as previously described ŽBlack et al., 1965.. For selected soil samples, the microbial population was estimated using a standard plate count ŽGhiorse and Blackwill, 1983.. Water samples were analyzed for pH, conductivity, total organic and inorganic carbon and total phosphate according to standard methods ŽAmerican Public Health Association, 1980.. Inorganic anions ŽCly, SO42y, . were analyzed by ion chromatography according to EPA method 300.0 ŽU.S. NOy 3 Environmental Protection Agency, 1984..

4. Results 4.1. Characterization of the test sites The physical and chemical characteristics of the subsurface soil and groundwater used to prepare the microcosms are summarized in Table 1. The solids from Texas, Oklahoma and the shallow sample from Louisiana were classified as sands, while the

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Table 1 Characterization of subsurface soil and groundwater Parameter

Louisiana site Shallow

Deep

Texas site

Oklahoma site Aerobic

Anaerobic

Subsurface soil Sampling depth Žm. Texture Sand Ž%. Silt Ž%. Clay Ž%. Organic carbon Ž%. Cation exchange capacity Žmeqr100 g. Standard plate count Žcellsrg.

0.6–0.9

3.4–4.6

7.0–7.6

0.5–1.0

1.8–2.4

95 2 3 0.2 2.8 –

38 46 16 0.8 10.0 –

84 10 6 0.3 14.3 1=10 5

93 1 6 0.2 6.9 3=10 4

95 1 4 0.1 6.9 –

Ground water Temperature Ž8C. pH Total organic carbon Žmgrl. Nonvolatile organic carbon Žmgrl. Total inorganic carbon Žmgrl. Conductivity Žmmhorcm. Total dissolved solids Žmgrl. Phosphate Žmgrl. Nitrate Žmgrl. Sulfate Žmgrl. Chloride Žmgrl.

21.9 7.6 23 1 24 440 286 - 0.1 2.6 38 33

27.0 6.8 381 100 233 37 000 – - 0.1 - 0.2 2200 28 200

14.4 7.6 48 – 132 1190 774 - 0.1 - 0.2 27 102

16.7 7.8 227 – 574 6220 4043 - 0.1 - 0.2 75 1012

deeper sample from the Louisiana site was classified as a silty clay. Organic carbon content for all of the soils was low, ranging from 0.1% to 0.8%. The microbial content of some samples was estimated using a standard plate count. Populations of aerobic heterotrophs were between 3 = 10 4 and 1 = 10 5 cellsrg, and are comparable to the numbers of culturable aerobic bacteria measured in sandy aquifers. Groundwater from the three sites varied considerably with respect to organic and inorganic constituents. Shallow water from the Louisiana site exhibited low levels of organic and inorganic carbon; the majority of the organic carbon was shown to be 1,2-dichloroethane. Concentrations of sulfate and chloride ion were 38 and 33 ppm, respectively. Groundwater from the Texas site exhibited high conductivity, which was consistent with the high concentrations several anions Žpossibly due to seawater intrusion.. Levels of chloride and sulfate ions detected in the water were 28 200 and 2200 mgrl, respectively. The sample also exhibited relatively high levels of organic carbon Ž381 mgrl., of which the majority was shown to be 1,2-dichloroethane. However, a portion Ž100 ppm. was shown to be nonvolatile, since it was not removed from the groundwater after purging for 12 h with nitrogen gas. The impact of landfill leachate on the groundwater from Oklahoma site was evident on the basis of differences in the levels of a number of constituents between the aerobic and anaerobic sampling sites. Elevated levels of total organic carbon, inorganic carbon, sulfate and inorganic chloride were present in samples collected downgradient of the landfill.

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4.2. Biotransformation of 1,2-dichloroethane in soil r water microcosms The potential for 1,2-dichloroethane biotransformation in aquifers with a history of exposure was examined under both anaerobic and aerobic conditions in soilrwater microcosms prepared using samples obtained from Louisiana site. Initial concentrations of the test chemical in the samples ranged from 41 to 47 ppm. Because the total nonvolatile organic carbon concentration of the groundwater was initially low, biotransformation under anaerobic conditions was examined in microcosms that had been amended with 500 ppm of sodium acetate. Throughout the investigation, the resazurin indicator present in the microcosms remained colorless, indicating a reduced state. As shown in Fig. 1, w14 Cx1,2-dichloroethane was readily degraded in microcosms prepared using the shallow sandy material; no lag period was evident prior to disappearance of the parent compound. After 126 days, over 75% of the parent compound was degraded in the active samples, whereas 95% was recovered in the controls. Analysis of isooctane extracts and the aqueous portion of the reaction mixtures failed to detect the presence of intermediate metabolites. However, w14 Cxethylene was detected in the headspace of the bottles by gas chromatography. Levels of w14 Cxethylene recovered in the samples after 126 days ranged from 45% to 63% of the initial radioactivity. The formation of w14 Cxethylene was attributed to biological activity, since it was not present in the killed controls. The accumulation of nonlabeled methane was

Fig. 1. Degradation of w14 Cx1,2-dichloroethane ŽDCA; 47 ppm. under anaerobic Žmethanogenic. conditions in Louisiana sandrwater microcosms Žclosed symbols. and in poisoned controls Žopen symbols.. Data points represent means of results from analysis of replicate microcosms; bars indicate the range of results for individual bottles.

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also detected in the gas phase of biologically active microcosms, indicating methanogenic conditions. The formation of 14 CO 2 was not detected in any of the samples. Similar results were obtained in microcosms prepared using the Louisiana silty clay Ždata not shown.. Levels of w14 Cxethylene detected in the microcosms after 56 days ranged from nondetectable to as high as 54% of the initial radioactivity. After 126 days, over 70% of the parent compound was degraded in biologically active microcosms. In contrast, biodegradation of w14 Cx1,2-dichloroethane was slow under aerobic conditions in microcosms prepared using the Louisiana sand. Throughout the study, the pink color of the redox indicator suggested that aerobic conditions were maintained in the microcosms. As shown in the Fig. 2, changes in w14 Cx1,2-dichloroethane concentrations were comparable in biologically active and killed controls during the first 66 days of incubation. However, after 94 days, evidence for aerobic biodegradation of the test chemical to 14 CO 2 was detected in several microcosms, suggesting variability in aerobic biodegradation potential. Levels of 14 CO 2 detected in several samples ranged from 3% to 43%, while negligible amounts of 14 CO 2 were detected in the killed controls. Anaerobic biotransformation of w14 Cx1,2-dichloroethane was subsequently examined in soilrwater microcosms prepared using soils from a contaminated site in Texas, as well as a site in Oklahoma with no known history of exposure. To evaluate the potential for biotransformation under naturally occurring conditions, the samples were not amended with an exogenous source of organic carbon. However, as previously noted, groundwater from both sites was shown to contain relatively high levels of dissolved organic carbon ŽTable 1..

Fig. 2. Degradation of w14 Cx1,2-dichloroethane ŽDCA; 41 ppm. under aerobic conditions in Louisiana sandrwater microcosms Žclosed symbols. and in poisoned controls Žopen symbols.. Data points represent means of results from analysis of replicate microcosms; bars indicate the range of results for individual bottles.

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As shown in Fig. 3, biotransformation of w14 Cx1,2-dichloroethane Žinitial concentration of 106 ppm. in microcosms prepared using aquifer material from the Texas site was initially slow. After 55 days, a visible change was observed in the microcosms, as the color of the soil changed from tan to black. It is likely that sulfate reducing conditions were established in the samples, due to the naturally high sulfate concentration of the groundwater. In addition, negligible levels of methane were detected in the headspace of the reaction mixtures. Following this change, w14 Cx1,2-dichloroethane biotransformation was apparent in the biologically active samples. After 168 days, 29% of the parent compound was biodegraded, based on the difference between observed losses of 43% and 14% in the active and control samples, respectively. Levels of w14 Cxethylene detected in the samples after 168 days were consistent with the stoichiometric conversion of 1,2-dichloroethane to ethylene. 1,2-Dichloroethane biotransformation was also observed following a 50-day lag period in microcosms prepared using aquifer material from the anaerobic site in Oklahoma ŽFig. 4.. The initial concentration of the test chemical in the microcosms was 29 ppm. It is likely that methanogenic conditions were established since the soil in the samples remained sandy brown in color throughout the study, and the fact that small amounts of methane were detected after 56 days during analysis of the headspace gases. However, since the groundwater used to prepare the microcosms contained sulfate ions Ž75 ppm., sulfate reducing conditions may also have existed in the samples. After 173 days, w14 Cx1,2-dichloroethane concentrations were reduced on the order of 20% to 64%

Fig. 3. Degradation of w14 Cx1,2-dichloroethane ŽDCA; 106 ppm. under anaerobic Žsulfate reducing. conditions in Texas soilrwater microcosms Žclosed symbols. and in poisoned controls Žopen symbols.. Data points represent means of results from analysis of replicate microcosms; bars indicate the range of results for individual bottles.

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Fig. 4. Degradation of w14 Cx1,2-dichloroethane ŽDCA; 29 ppm. under anaerobic Žmethanogenic. conditions in Oklahoma soilrwater microcosms Žclosed symbols. and in poisoned controls Žopen symbols.. Data points represent means of results from analysis of replicate microcosms; bars indicate the range of results for individual bottles.

in the biologically active samples as compared to losses of only 5% in the controls. w14 Cxethylene was the only metabolite detected and the samples. At the completion of the study, levels of w14 Cxethylene detected in the microcosms ranged from 38% to 48% of the initial radioactivity. Additional microcosms were prepared using aquifer materials from both Texas and Oklahoma to examine potential for biodegradation under aerobic conditions. No degradation was apparent under aerobic conditions Žas evident from the pink color of the redox indicator., since changes in 1,2-dichloroethane concentrations in biologically active microcosms were comparable to the controls, and negligible levels of 14 CO 2 were detected following 180 days of incubation Ždata not shown.. 4.3. Effects of organic cosubstrates on anaerobic biotransformation of 1,2-dichloroethane The effects of various organic amendments on the anaerobic biotransformation of 1,2-dichloroethane was examined in microcosms prepared using the silty-clay from the Louisiana site. Initial 1,2-dichloroethane concentrations were approximately 50 ppm Ž0.5 mM., and all microcosms were amended with nominal 5.0 mM levels of the organic amendments. As summarized in Table 2, methanogenic conditions, as indicated by gas production, and the reductive dehalogenation of 1,2-dichloroethane to ethylene were observed within 7 weeks in microcosms amended with a variety of organic amendments.

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Table 2 Effects of organic cosubstrates on the anaerobic biotransformation of 1,2-dichloroethane in Louisiana sandrwater microcosms Gas production Žml. and 14 Cx1,2-dichloroethane degradation Ž% DCA loss. with time Amendment

Control Acetate Butyrate Lactate Ethanol Propylene glycol Glycerol Glucose

7 weeks

14 weeks

Gas

% DCA loss

Gas

% DCA loss

0.2 0.6 0.3 0.8 0.2 0.7 0.4 3.5

11.2 53.3 46.4 63.5 67.7 71.9 42.9 23.9

0.1 0.7 0.8 0.7 0.2 0.9 0.9 2.9

4.8 69.4 68.9 50.6 70.7 67.2 71.3 56.7

After 14 weeks, 1,2-dichloroethane concentrations were reduced in all biologically active samples by 57% to 71% as compared to the nonamended control.

5. Discussion Reductive dehalogenation appeared to be the dominant mechanism involved in the biotransformation of 1,2-dichloroethane under anaerobic conditions. Reductive dehalogenation refers to an oxidation–reduction reaction in which electrons are transferred from a donor Že.g., reduced organic substrates. to the chlorinated hydrocarbon acceptor, resulting in displacement of a chlorine substituent ŽVogel et al., 1987.. When a chlorinated alkane contains two chlorine substituents on adjacent carbon atoms, reductive dihaloelimination is frequently observed, resulting in the formation of an alkene ŽVogel et al., 1987.. The anaerobic biotransformation of 1,2-dichloroethane to ethene in aquifer material is consistent with the work of others with pure cultures of methanogenic and sulfate reducing bacteria ŽBelay and Daniels, 1987; Egli et al., 1987.. Similarly, dihaloelimination reactions have been reported for the biotransformation of hexachloroethane to tetrachloroethene in aquifer material ŽCriddle et al., 1986., and for a series of halogenated ethanes Žvicinal halides. in anoxic sediments ŽJafvert and Wolfe, 1987.. In contrast, Holliger et al. Ž1990a. reported the biotransformation of 1,2-dichloroethane by methanogenic bacteria proceeds via both dihaloelimination and hydrogenolysis reactions. Multiple pathways have also been reported for the biotransformation of 1,1,2,2-tetrachloroethane in anaerobic digester sludge, although dihaloelimination was the dominant reaction, accounting for approximately 95% of the observed reaction ŽChen et al., 1996.. Thus, although experiments with bacterial cultures indicate the potential for multiple pathways exists, results obtained with environmental samples suggest that dihaloelimation would be the dominant reaction in strongly reducing groundwater environments. Microbial adaptation appeared to be required for the anaerobic biotransformation of 1,2-dichloroethane. Although no lag period was evident in microcosms prepared from

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the Louisiana site, lag periods ranging from 7 to 8 weeks were noted in microcosms prepared using the Texas and Oklahoma materials. Differences in lag periods between the samples may have been due to presence or absence of readily utilizable organic nutrients. However, the absence of a lag period in the Louisiana soil is consistent with evidence for natural biotransformation in situ based on correlations between organic carbon and inorganic chloride levels in the groundwater ŽKlecka ˇ et al., 1994.. In contrast, Wood et al. Ž1981. and Jafvert and Wolfe Ž1987. reported negligible biotransformation of 1,2-dichloroethane in anaerobic sediment–water microcosms prepared using sediments from surface environments. Since in each case, the sediments were shown to be capable of reductive dehalogenation of other halogenated aliphatic compounds, it is unlikely that systems were limited by organic carbon or other electron donor. It is possible that the short incubation periods Ž30 to 35 days. used in the latter studies did not allow sufficient time for microbial adaptation to occur. The fact that anaerobic biotransformation was observed in aquifer materials collected from sites with diverse physical and chemical characteristics and exposure histories suggests that the potential for degradation may be relatively widespread in the environment. A variety of anaerobic microorganisms with the ability to degrade 1,2-dichloroethane have been described, including methanogens and sulfate-reducing bacteria ŽBelay and Daniels, 1987; Egli et al., 1987; Holliger et al., 1990a.. In addition, Holliger et al. Ž1990b. reported the degradation of 1,2-dichloroethane by cell extracts of methanogenic bacteria correlated with presence of enzyme cofactors, including vitamin B12 and factor F430. Schanke and Wackett Ž1992. have also noted that many microbial reductive dechlorination reactions are mimicked in laboratory experiments with transition metal coenzymes, and that these biological cofactors are often found at high concentrations in anaerobic bacteria. Only limited degradation was observed in the present study under aerobic conditions, in contrast to the variety of 1,2-dichloroethane degrading bacteria that have been described ŽStucki et al., 1983; Janssen et al., 1985; Stucki et al., 1987; Vandenbergh and Kunka, 1988; van den Wijngaard et al., 1992.. In addition, Watwood et al. Ž1991. reported mineralization of 1,2-dichloroethane in soilrwater microcosms prepared with surface soils Ža sandy aridisol and a riparian riverbank soil. from New Mexico. The limited aerobic degradation in the present study may be due to differences in the environmental conditions and exposure histories among the sampling sites. Alternatively, the extended lag periods may reflect changes that are required for the evolution of 1,2-dichloroethane degrading organisms. Janssen et al. Ž1995. have speculated that while several enzymes necessary for aerobic biodegradation appear to be ubiquitous in nature, two key enzymes appear to have evolved specifically for the degradation of 1,2-dichloroethane. Since microbial adaptation to specific contaminants has been proposed to occur as a result of the contamination ŽNishino et al., 1994., the environmental conditions and exposure history are likely to play an important role in the development of aerobic 1,2-dichloroethane degrading bacteria in subsurface environments. The ability of naturally occurring microorganisms to degrade 1,2-dichloroethane has potential application in the remediation of environmental contamination. For example, Michiels and Breugelmans Ž1993. have recently described results of a field application of enhanced aerobic bioremediation for the in situ treatment of an aquifer contaminated

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with 1,2-dichloroethane. Anaerobic methods are also being developed for bioremediation of chlorinated hydrocarbons in groundwater and results of various field trials have been described ŽBoyer et al., 1988; Semprini et al., 1991; Beeman et al., 1994.. Based on results of the present investigation, anaerobic bioremediation was proposed for remediation of 1,2-dichloroethane contamination at the Louisiana site ŽKlecka ˇ et al., 1994., however, the low permeability of the contaminated zone precluded implementation at the site. As an alternative to enhanced bioremediation technologies, there has been increasing recognition of the fact that natural processes alone may be sufficient to mitigate the impact of groundwater contamination on human health and the environment. Over the past several years, natural attenuation has become increasingly accepted as a remedial alternative for organic contaminants in groundwater. The potential for natural attenuation of 1,2-dichloroethane has been examined by several investigators ŽLee et al., 1996; Bosma et al., 1997.. Bosma et al. Ž1997. described results of an investigation conducted at a vinyl chloride manufacturing site where 1,2-dichloroethane was the primary groundwater contaminant. Based on changes in 1,2-dichloroethane concentrations along the groundwater flow path, the half-life for degradation under field conditions was estimated to be in the range from less than 1 year to over 30 years. Lee et al. Ž1996. have also presented results of an investigation of the intrinsic bioremediation of 1,2-dichloroethane at a Gulf Coast manufacturing site. Degradation products detected in the groundwater included 2-chloroethanol, ethanol, ethene and ethane, suggesting both aerobic and anaerobic biodegradation. Half-lives for estimated for 1,2-dichloroethane degradation under field conditions ranged from 64 to 165 days. The degradation of chlorinated hydrocarbons in groundwater has attracted considerable attention over the last decade. The present study provides additional information that can be used to assess the fate and lifetime of 1,2-dichloroethane in the environment. In view of the reaction rates, biological transformations are likely to play a major role in determining the fate of the compound in both aerobic and anaerobic environments. Because the conditions in groundwater vary considerably, an understanding of the physical, chemical and biological characteristics is essential to predicting the relative importance of abiotic and biological processes in determining the fate and lifetime of 1,2-dichloroethane in subsurface environments. Acknowledgements The authors wish to thank Pat Broussard, Charlie Cooper and Jim Presley of EDC Engineering, John Davis, Jill McCullough and Glyn Boudreaux of Dow Chemical, and Neil Adrian and Joseph Suflita of the University of Oklahoma for assistance in obtaining the aquifer materials. References Adrian, N.R., Robinson, J.A., Suflita, J.M., 1994. Spatial variability in biodegradation rates as evidenced by methane production from an aquifer. Appl. Environ. Microbiol. 60, 3632–3639.

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