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Environment & Resources, Bygningstorvet-Building 115, Technical University of Denmark, DK-2800 Lyngby, Denmark
* Corresponding author (chs{at}er.dtu.dk).
Received for publication November 13, 2002.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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Abbreviations: CFC, chlorofluorocarbon • HCFC, hydrochlorofluorocarbon • HFC, hydrofluorocarbon • LFG, landfill gas • MMO, methane monooxygenase • VOC, volatile organic compound
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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In several cases LFG has been shown to contain more than 100 different trace gases including halogenated and aromatic hydrocarbons and sulfur- and oxygen-containing compounds (Rettenberger and Stegmann, 1996; Eklund et al., 1998; Allen et al., 1997). Typical trace gas concentrations are in the range of 10 to 250 mg m–3. The trace components originate from hazardous materials deposited in the landfill or from biological or chemical degradation of materials disposed of in the landfill. Due to pressure and concentration gradients, the LFG will migrate through the surrounding soil of landfills, causing emission of LFG into the atmosphere.
Aromatic and chlorinated hydrocarbons are widely used in industry as solvents and degreasing agents. Fluorinated hydrocarbons have been used in a number of industrial processes and products like refrigerating aggregates, foaming agents, solvents, and propellants. Landfills receiving industrial wastes might be expected to have a high content of these compounds. However, an investigation of 23 American landfills receiving household waste showed that in 85% of these landfills benzene and VC were found in the LFG (Brosseau and Heitz, 1994). Emission of trace gases is a potential risk to human health, as compounds like VC and benzene are proven carcinogens (Christensen and Kjeldsen, 1995). Emission of trace components like CFCs is of global environmental concern due to their involvement in the depletion of the ozone layer and global warming (Wallington et al., 1994).
Microbial oxidation of methane in aerobic soils plays a significant role in reducing the emission of methane to the atmosphere (Lelieveld et al., 1998). Aerobic soils serve as biofilters for methane produced in anoxic soils, and some soils even serve as sinks for atmospheric methane (Oremland and Culbertson, 1992). Soils exposed to elevated concentrations of methane can develop a high capacity for methane oxidation by selection of methanotrophic bacteria. In landfill covers methane and oxygen countergradients may appear due to emission of methane from the waste and diffusion of oxygen from ambient air. Oxidation of methane in landfill top cover soil has been shown to reduce the amount of methane emitted to the atmosphere (Czepiel et al., 1996; Liptay et al., 1998; Bogner et al., 1997; Christophersen et al., 2001). A defining characteristic of the methanotrophic bacteria is the enzyme methane monooxygenase (MMO), which catalyzes the oxidation of methane. While all methanotrophs can synthesize a particulate or membrane-bound form (pMMO) of the enzyme, some can also produce a soluble form (sMMO), which has very broad substrate specificity and is known to rapidly cometabolize a variety of aliphatic compounds, including some halogenated hydrocarbons (Hanson and Hanson, 1996). It is therefore reasonable to believe that trace components might be cometabolically degraded during migration in landfill cover soils. Kjeldsen et al. (1997) published results on oxidation of methane and aromatic hydrocarbons (benzene and toluene) and cometabolic degradation of TCE and 1,1,1-trichloroethane (TCA) in soil incubated with LFG. High methane oxidation potentials and high degradation rates for benzene and toluene were found. In addition, slow cometabolic degradation of TCE and 1,1,1-TCA was observed in the presence of methane. The authors concluded that degradation processes might have a significant effect on the emission from landfill covers of the compounds studied. These results provide the potential for further investigation of the topic.
The objective of the current study was to investigate the potential for natural attenuation of methane and VOCs in soil exposed to LFG. Maximal oxidation rates were determined in batch experiments where soil was incubated with methane and selected VOCs. The degradability of 26 trace components was investigated, including chlorinated methanes, ethanes, ethenes, fluorinated hydrocarbons, and aromatic hydrocarbons. The extent of the zone with methanotrophic activity was determined by incubation of soil sampled from different depths at a location emitting methane. Oxidation capacities for landfill soil covers were calculated using the obtained oxidation rates.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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Soil texture was determined by sieve analyses and classified according to the USDA classification system (Soil Survey Staff, 1975). The field bulk density was determined by weighing soil-filled steel cylinders sampled in the field, subtracting the cylinder weight, and dividing by the cylinder volume. Soil moisture content was determined gravimetrically by oven-drying at 105°C for 24 h and expressed as the mass ratio of water to dry soil. Soil organic content was determined by loss on ignition. The pH was measured in soil water suspensions (10 g soil to 25 mL 0.01 M CaCl2 solution). Ammonium was measured by extracting 50 g soil with 100 mL 0.1 M KCl for 1 h followed by filtration through a 0.45-µm filter. Ammonium concentrations in the filtrates were determined colorimetrically with an autoanalyzer (AutoAnalyzerII, Technicon, Tarrytown, NY). Anion concentrations (NO–3 and NO–2) were analyzed by extracting 20 g soil with 30 mL Milli Q water (Millipore, Billerica, MA) for 1 h followed by filtration through a 0.45-µm filter. Anion concentrations were analyzed by ion chromatography (Model DX120; Dionex, Sunnyvale, CA). The copper concentration was determined after soil extraction with Titriplex III (Merck, Darmstadt, Germany) by atomic adsorption spectrometry. All soil concentrations are expressed as mass of dry soil.
Soil Microcosms
The oxidation process was examined in glass bottles (117 mL in total volume) equipped with Mininert valves made of Teflon (VICI AG, Schenkon, Switzerland). The valves enabled gas to be sampled or injected by a hypodermic needle and a syringe. A fixed amount of soil (20 g moist soil) was added to each batch container. To obtain methane oxidation conditions, air was withdrawn from each container using a syringe and replaced with methane and oxygen, which gave an initial mixture of methane (15% v/v), oxygen (35% v/v), and nitrogen (50% v/v). Gas samples containing the test compound were removed from gaseous stock solutions by a gas-tight glass syringe and injected into the batch containers. The degradation of the VOCs was studied in single compound tests. The initial concentrations used were generally in the range of typical trace gas concentrations in LFG (10–250 mg m–3). However, higher initial concentrations for some of the lower or nonhalogenated compounds were necessary for better analytical sensitivity using chromatographic analysis. Gas samples withdrawn from the headspace were sampled periodically and analyzed by gas chromatography. Between sampling, the bottles were turned to ensure total mixed conditions in the batches. The batch experiments were conducted at room temperature (22°C). All batch experiments were performed in four replicates.
To control for nonmicrobial processes (abiotic degradation, sorption, or volatilization), sodium azide (25 mg kg–1 wet soil) was added to avoid microbial growth in control batches.
In general, the batch experiments were all conducted with soil sampled at 15 to 20 cm below the surface. Furthermore, soil samples from successive depth intervals were collected to determine the distribution of oxidation capacity in the soil profile. The soil was sampled in 5-cm intervals from the surface to a 30-cm depth and in 10-cm intervals from 30 to 90 cm below the surface. Soil samples were incubated with methane and selected trace components including the hydrochlorofluorocarbons (HCFCs) HCFC-21 and HCFC-22, VC, benzene, and toluene.
Chemicals
PCE, TCE, TeCM, TCM, DCM, benzene, toluene, ethylbenzene and o-xylene were obtained from Merck. Cis-1,2-DCE, trans-1,2-DCE, and 1,1-DCE were obtained from Fluka (Buchs, Switzerland). The chlorinated ethanes (1,1,2,2-TeCA, 1,1,1-TCA, 1,1,2-TCA, 1,2-DCA, and 1,1-DCA) were purchased from Aldrich (Steinheim, Germany). CFC-11, CFC-12, HCFC-21, and HCFC-22 were purchased from Flourochem Limited (Old Glossop, England). Hydrofluorocarbons (HFCs) HFC-134a and HFC-245fa were obtained from Interchim (Montluçon, France). HCFC-141b was obtained from Honeywell (Weert, the Netherlands). Vinyl chloride was obtained from Gerling Holz & Co. (Hamburg, Germany). All solvents were obtained in high purity. Full chemical names, together with selected physicochemical parameters of all compounds, are listed in Table 1.
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The main gas components (CH4, CO2, O2, and N2) were analyzed on a Chrompack Micro GC CP-2002P gas chromatograph (Chrompack) equipped with a thermal conductivity detector and two columns. Oxygen and nitrogen were quantified on a 4-m-long Molsieve 5A column (Chrompack) and methane and carbon dioxide on a 10-m-long Poraplot Q column (Chrompack). The carrier gas was helium and the column temperature was 40°C. Gas standards produced by MicroLab (Aarhus, Denmark) ranging from 0.02 to 50% v/v were used for calibration.
Data Evaluation
From the measured gas concentrations, the total amount (µg) of test compound in each batch was determined by phase distribution calculations. Assuming steady state, the total amount of a chemical compound (MTotal) can be expressed as a function of the measured gas concentration (Cg) using Henry's law and the soil–water distribution coefficient Kd based on the octanol–water distribution coefficient Koc:
 | [1] |
The physicochemical parameters used are given in Table 1. For the main components (CH4, O2, CO2, and N2) only distribution between water and air was considered. Control experiments with chemically sterilized soil indicated a rapid adjustment of the equilibrium between phases. A slow decrease in the gas phase could indicate loss due to sorption or dissolution. El-Farhan et al. (2000) investigated the kinetics of TCE cometabolism and toluene consumption in similar soil batch systems and found that the kinetic parameters describing the degradation were not affected by the physical processes like sorption and dissolution taking place in the soil system.
The kinetics of oxidation were examined by plotting the total concentration of the compound versus time. In all cases the methane oxidation followed zero-order kinetics and maximal oxidation rates were calculated by applying zero-order kinetics to the data. However, in some cases the decrease of the VOC concentrations at the end of the experiments limited the reaction and changed it into first order, resulting in a decreasing oxidation rate. To compare the degradation rates between different compounds, maximal rates were calculated applying zero-order kinetics to the data covering the period from the start of the experiment and until the degradation was 90% complete (C/C0 = 10%). The zero-order rate constant was normalized to the dry soil mass. For each compound, an average oxidation rate was calculated based on four replicates. In this paper the term "degradation" is used when a decrease in concentration was observed over time in microbially active experiments.
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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Methane Oxidation and Degradation of Trace Components in Soil Microcosms
In general, good reproducibility was obtained and results from four replicate batches were almost identical. To examine reproducibility, 10 replicates were incubated with a mixture of methane, HCFC-21, and HCFC-22. The average methane oxidation rate was 82 µg g–1 h–1 with a standard deviation of 2.4. The average oxidation rates for HCFC-21 and HCFC-22 were 0.51 and 0.29 µg g–1 h–1 with standard deviations of 0.02 and 0.01, respectively. Figure 1
shows the degradation of HCFC-21 and HCFC-22 in 5 out of the 10 incubation bottles. Figure 1 also shows the results from two corresponding control experiments.
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Degradation of Chlorinated Hydrocarbons in Soil Microcosms
Both TCM and DCM were degraded in the presence of oxygen and methane and the degradation occurred in parallel with the oxidation of methane (Fig. 3A)
. However, degradation of TeCM was not observed. Maximal oxidation rates, initial concentrations, and regression coefficients are listed in Table 1. The degradation of DCM occurred 20 times faster than TCM, which is consistent with the general tendency that degradation of chlorinated aliphatic compounds is inversely related to the chlorine to carbon ratios under aerobic conditions.
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Kjeldsen et al. (1997) performed similar experiments with soil from Skellingsted Landfill and found much lower oxidation rates for both methane and TCE (3.75 and 0.00125 µg g–1 h–1, respectively). This can be explained by a combination of several factors. Kjeldsen et al. (1997) collected soil from other sampling sites and had less favorable experimental conditions, including lower temperatures (10°C), deeper sampling depths (50–60 cm below the surface), and soil dehydration during storage.
Transformation yield has been used to describe cometabolic rates, and is defined as the amount of compound transformed per amount of primary substrate utilized (Alvarez-Cohen and McCarty, 1991a). The transformation yields are often quite consistent, ranging from 15 to 50 µg TCE mg–1 methane for a variety of cultures (Alvarez-Cohen and Speitel, 2001). In general the transformation yields in this study were rather low: for example, the transformation yield for TCE was 1.2 µg mg–1 methane. The lower transformation yields may be a result of the fact that not all the methane oxidizers present in the soil are capable of cometabolizing VOCs.
In general, fully chlorinated aliphatics are expected to persist under aerobic conditions since these compounds do not support growth for aerobic microorganisms (Janssen and Koning, 1995). Especially, PCE and TeCM are generally considered to be resistant to aerobic cometabolic degradation by methanotrophs (Oldenhuis et al., 1989).
Both 1,1-DCA and 1,2-DCA were degraded; 1,2-DCA much more rapidly than 1,1-DCA, with maximal oxidation rates comparable with trans-1,2-DCE and VC (Table 1 and Fig. 3C). Also 1,1,2-TCA was degraded with a rate comparable with 1,1-DCE. Neither 1,1,1-TCA nor 1,1,2,2-TeCA were degraded within the time duration of the experiment (Fig. 3C). The results agree well with those obtained by Henson et al. (1989), who observed that 1,2-DCA was more rapidly transformed than 1,1-DCA and 1,1,2-TCA, while 1,1,1-TCA was not removed in a mixed bacterial culture grown on methane. Literature surveys on degradation of 1,1,1-TCA by methanotrophs show some inconsistency. Strand et al. (1990) reported slow degradation of 1,1,1-TCA in batch experiments with mixed cultures of methanotrophs incubated with methane. Likewise, Oldenhuis et al. (1989) observed transformation but not complete degradation of 1,1,1-TCA by M. trichosporium OB3b grown in continuous cultures. Methylosinus trichosporium OB3b PP358, which constitutively expresses soluble MMO, did not degrade 1,1,1-TCA, even though it rapidly degraded a number of other chlorinated solvents (Aziz et al., 1999).
Degradation of Fluorinated Hydrocarbons in Soil Microcosms
Of the HCFCs only HCFC-21 and HCFC-22 were shown to be degradable, while HCFC-141b was not (Fig. 3E). The compound HCFC-21 was more rapidly degraded than HCFC-22, which had a degradation rate closer to TCM (Table 1). The compounds CFC-11, CFC-12, CFC-113, HFC-134a, HCFC-141b, and HFC-245fa did not seem to be degradable within the duration of the experiment (Fig. 3D, 3E, 3F). Information on cometabolic degradation of fluorinated hydrocarbons reported in the open literature is very limited. DeFlaun et al. (1992) studied the oxidation of HCFCs by M. trichosporium OB3b and obtained aerobic degradation of three out of the five HCFCs tested, including HCFC-21 and HCFC-141b. Complete dehalogenation of HCFC-21 was verified by ion chromatographic analysis of the stoichiometric release of chloride and fluoride. However, CFC-11 and HFC-134a were not degraded. Streger et al. (1999) studied the degradation of hydrohalocarbons by three strains of naturally occurring methanotrophs. All three strains completely transformed HCFC-21. Ion chromatography analysis showed a release of stoichiometric amounts of chloride. One of the bacterial strains, M. trichosporium OB3b, also transformed HCFC-141b. None of the tested bacteria were able to degrade HFC-134a. Chang and Criddle (1995) report biodegradation of both HFC-134a and HCFC-22 in a mixed methanotrophic culture; HCFC-22 was more rapidly transformed (15 times) compared with HFC-134a. The mixed methanotrophic culture also contained at least three heterotrophs, and it is therefore possible that they facilitated the degradation of, for example, HFC-134a. The negative impact of CFC and HCFCs on the stratospheric ozone layer has prompted an effort for environmentally acceptable alternatives like the nonchlorinated HFCs. The compound HFC-245fa is a new substitute for HCFCs used as blowing agents in insulation foams and no references in the open literature on cometabolic degradation by methanotrophs are known to the authors. Chlorofluorocarbons are considered inert toward biodegradation under aerobic conditions (Key et al., 1997). The presence of one or more hydrogen atoms in HCFCs and HFCs makes these compounds susceptible to undergo oxidation. The strong chemical carbon–fluoride bond provides greater stability to the HFCs compared with chlorinated hydrocarbons, suggesting that HFCs in general would be more resistant to microbial degradation. Although HFCs are good replacements for HCFCs as a result of their zero ozone depletion potential, their contribution to global warming may be important if they are resistant to microbial transformation.
Degradation of Aromatic Hydrocarbons in Soil Microcosms
The aromatic hydrocarbons were rapidly degraded, giving high maximal oxidation rates between 0.17 and 1.4 µg g–1 h–1 (Table 1 and Fig. 3G). The aerobic biological degradation of benzene, toluene, ethylbenzene, and xylene (BTEX) is rather well understood, and several of the compounds have previously been shown to be degradable under aerobic conditions. A large number of different bacteria are known to degrade benzene and toluene; some are also able to convert chlorinated aliphatics by cometabolism with benzene or toluene as primary substrates (Rivas and Arvin, 2000; Lu et al., 1998; Landa et al., 1994). Aromatic hydrocarbons are often detected in LFG due to their widespread utilization and high persistence under anaerobic conditions. Emission of aromatics from landfills is of special public concern since benzene is a proven carcinogenic. In general, the sterilized control experiments showed no decrease in the VOC concentrations, indicating that microbial oxidation was the only explanation for the decrease in the active experiments.
Depth Distribution of Oxidation Activity
Soil samples were collected at the test station from different depths to determine the vertical distribution of methane oxidation and degradation of selected trace components. The methane oxidizers were very active in oxidizing methane and trace components down to a depth of 50 cm below the surface (Fig. 4A, 4B)
. Maximum rates were obtained with soil from 15 to 20 cm. Both Kightley et al. (1995) and De Visscher et al. (1999) observed peak oxidation in soil cores incubated with methane between 10 and 20 cm deep where both oxygen and methane were present. Christophersen and Kjeldsen (2001) performed an extensive field study at Skellingsted landfill, in which vertical soil profiles measured every second week during a one-year period showed that both methane and oxygen often were present between 20 and 40 cm deep. This agrees with the results from this study since the maximum oxidation zone is expected to form in a soil layer of overlapping O2 and CH4 gradients. The oxidation rates decreased dramatically at the 50-cm depth as a result of O2 limitation, which is supported by soil air measurements conducted at the 60-cm depth showing a gas composition of mainly methane and carbon dioxide and very low or no oxygen presence (Fig. 4C). The low methane oxidation rate (3.75 µg g–1 h–1) obtained by Kjeldsen et al. (1997) also with soil from Skellingsted landfill is most likely due to the deeper sampling between 50 to 60 cm below the surface. The distribution of methane oxidizers in a relatively narrow zone makes soil sampling depth a critical parameter when measuring oxidation capacities of landfill soil covers in laboratory incubation experiments.
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PERSPECTIVES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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b is the soil bulk density (Mg m–3). The flow of LFG is considered intergranular with no macropore flow due to worm holes, cracks, and fissures, which will make the flow conditions very inhomogeneous. The maximum concentration, Cmax (g m–3), of the LFG compound that can be degraded while passing through the oxidative zone is then:
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For the lower chlorinated hydrocarbons (DCM, DCEs, DCAs, HCFC-21, and HCFC-22), the maximal degradable concentrations by far exceed the maximal measured concentrations in LFG, which indicates that these compounds may be totally degraded in the top soil cover. The degradation rates are lower for the compounds with higher chlorine substitution, thus resulting in smaller maximal degradable concentrations for these compounds. Nevertheless, for higher chlorinated compounds like TCM and TCE, the maximum degradable concentrations by far exceed the maximum measured concentrations in LFG, which indicates that these compounds also may be totally degraded in the soil cover.
Degradation of fully substituted carbons (TeCM, PCE, CFC-11, CFC-12, and CFC-113) in the oxidative zone is limited, but these compounds have been found degradable under anaerobic conditions. Under methanogenic conditions, which often exist in the lower part of soil covers and within the waste, CFC-11 and CFC-12 may undergo reductive dehalogenation leading to accumulation of lesser chlorinated compounds, like products HCFC-21 and HCFC-22 (Ejlertsson et al., 1996; Deipser and Stegmann, 1997), which then might be degraded in the oxidative zone in the surrounding soil. Likewise, the anaerobic sequential halogen–hydrogen substitution of PCE will generate the far more toxic VC. Vinyl chloride will be rapidly transformed under aerobic conditions in, for example, the soil covers. The same could be valid for other highly halogenated organic compounds, which are more restricted to degradation in the aerobic zone, like trichloroethanes and carbon tetrachloride.
The above evaluation is based on high oxidation rates obtained under optimal conditions and assumes a homogeneous soil layer with an intergranular flow of LFG. However, landfill top covers are highly dynamic systems influenced by atmospheric pressure, temperature, and precipitation. During cold periods with low temperatures the emission might increase due to lower microbial activity. Decreasing barometric pressure also causes higher emissions by enhancing the advective gas flow out of the landfill and thus shrinking the oxic zone. During warm periods with low precipitation the soil may dry out and limit the activity of the methanotrophic bacteria and thereby increase the methane emission.
However, this study demonstrates that LFG-affected soil shows a significant potential for methane oxidation and biodegradation of VOCs. At old landfills with lower gas production, methane oxidation and degradation of VOCs in cover soils may play a very important role in reducing the emission of both methane and trace components into the atmosphere.
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REFERENCES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONPERSPECTIVESREFERENCES |
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