Journal of Environmental Quality 30:369-376 (2001)
© 2001 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
TECHNICAL REPORT
ATMOSPHERIC POLLUTANTS AND TRACE GASES
Methane Oxidation in Two Swedish Landfill Covers Measured with Carbon-13 to Carbon-12 Isotope Ratios
Gunnar Börjessona,
Jeffrey Chantonb and
Bo H. Svenssona
a Dep. of Water and Environmental Studies, Linköping Univ., SE-581 83 Linköping, Sweden
b Dep. of Oceanography, Florida State Univ., Tallahassee, FL 32306-4320
Corresponding author (Gunnar.Borjesson{at}tema.liu.se)
Received for publication November 12, 1999.
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ABSTRACT
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The release of methane (CH4) from landfills to the atmosphere and the oxidation of CH4 in the cover soils were quantified with static chambers and a 13C-isotope technique on two landfills in Sweden. One of the landfills had been closed and covered 17 years before this investigation while the other was recently covered. On both landfills, the tops of the landfills were compared with the sloping parts in the summer and winter. Emitted CH4, captured in chambers, was significantly enriched in 13C during summer compared with winter (P < 0.0001), and was enriched relative to anaerobic-zone methane. The difference between emitted and anaerobic zone 
13C–CH4 was used to estimate soil methane oxidation. In summer, these differences ranged from 9 to 26
, and CH4 oxidation was estimated to be between 41 and 50% of the produced CH4 in the new landfill, and between 60 and 94% in the old landfill. In winter, when soil temperature was below 0°C, no difference in 13C was observed between emitted and anaerobic-zone CH4, suggesting that there was no soil oxidation. The temperature effect shown in this experiment suggests that there may be both seasonal and latitudinal differences in the importance of landfill CH4 oxidation. Finally the isotopic fractionation factor (
) varied from 1.023 to 1.038 and was temperature dependent, increasing at colder temperatures. Methanotrophic bacteria appeared to have high growth efficiencies and the majority of the methane consumed in incubations did not result in immediate CO2 production.
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INTRODUCTION
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AMONG greenhouse gases, CH4 is one of the most important, and its atmospheric concentration is increasing about 0.6% yr-1 due to anthropogenic activities (Intergovernmental Panel on Climate Change, 1996). Global CH4 emissions from landfills have been estimated at 40 Tg yr-1 (Prather et al., 1994), but with a broad range of uncertainty (20–70 Tg yr-1). This is more than 10% of all the anthropogenic sources (375 Tg), and landfilling is thereby the largest individual anthropogenic source of CH4 in many countries (Intergovernmental Panel on Climate Change, 1996). Most of this landfill CH4 is derived from the industrialized parts of the world (Subak et al., 1993), which are situated in temperate climate zones.
Methane in landfills is produced through anaerobic microbial degradation of organic matter under constant temperature conditions, despite variations in surrounding atmospheric temperature. The temperature-induced variation in the CH4 emission rate from landfills, with less CH4 emission during warm periods and greater emissions during cold periods, is controlled by the oxidation of CH4 in the aerobic surface (Börjesson and Svensson, 1997a; Chanton and Liptay, 2000).
In order to estimate the ratio of CH4 oxidized to CH4 produced in landfills, the use of 13C–CH4 analysis has been shown to be one of the most useful methods (Liptay et al., 1998; Bergamaschi et al., 1998; Chanton and Liptay, 2000). Methane-oxidizing microorganisms show a slight discrimination against methane containing 13C. Methane will therefore get heavier (13C-enriched) on its way through landfill cover soils when these microorganisms are active. Results obtained with this method suggest ratios between oxidized and produced CH4 in landfills at around 10% in the northeastern USA (Liptay et al., 1998), 25 to 35% in Florida, USA (Chanton and Liptay, 2000), and 39 to 53% in Germany (Bergamaschi et al., 1998). Obviously, more data are needed for both warmer and colder climates. The aim of this investigation was to study the effects of different temperatures on the oxidation process with the use of 13C analysis.
At subsites where CH4 leaks out, CH4 oxidation activity is often correspondingly high (Adamse et al., 1972; Jones and Nedwell, 1993; Börjesson et al., 1998b). On landfills, the sloping sides often show high CH4 emissions, since they usually are not well covered, due to erosion or difficulties in the application of cover material. Therefore, soil CH4 oxidation was studied and compared in both the sloping sides and tops of landfills. To our knowledge, such a contrast has not been made.
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MATERIALS AND METHODS
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Sampling Sites
Two Swedish landfill sites were chosen for the experiment: a recently closed (May 1997) landfill near Falköping and an older landfill in Hökhuvud that was closed in 1980. Falköping has a gas extraction system with vertical wells, and landfill gas is converted to heat and distributed to a neighboring dairy. Hökhuvud has no gas recovery system. Both are filled with municipal solid waste from rather small communities (25000–35000 inhabitants). More details on the landfills and their cover soils are given in Table 1.
Field Measurements
Both sites were divided into two parts: top and slope. At least eight static chambers (7.4-L cans, base area 0.0369 m2) were placed randomly on each area in August 1997 and January or February 1998. In addition to this schedule, in March 1998 gas samples were also collected from chambers placed on prescanned (with a Gas-trac NGX-8 [Northern Illinois Gas Company, Naperville, IL]; Börjesson and Svensson, 1997a) spots where CH4 concentrations were obviously elevated, indicating fissures or a thin layer of cover soil. From each static chamber gas samples were taken every 5 min over a 20 min interval, with the use of pre-evacuated 10-mL glass vials connected to the can with a double needle through a rubber membrane in the chamber (Börjesson and Svensson, 1997a). From the static chambers, gas samples were also withdrawn for 13C analyses and stored in glass vials with butyl rubber stoppers. Gas samples were analyzed for CH4 and CO2 on a Chrompack (Bergen-op-Zoom, the Netherlands) CP9001, with a flame ionization detector following a methanizer (Börjesson and Svensson, 1997b).
Gas samples were also taken from the anaerobic part of the landfills. At Hökhuvud, gas samples from different depths (down to 1.2 m) were taken with a probe inserted into 22-mm auger holes (equipment described by Börjesson and Svensson, 1997a), while in Falköping gas samples of the anaerobic zone were taken from the gas collection system. These samples were transferred to pre-evacuated 100-mL glass flasks for later analysis of 13C content. Samples were also withdrawn and transferred to argon-flushed 25-mL borosilicate glass tubes, for later gas-chromatographic analysis of CH4, N2, and O2 (Carlo Erba [Milan, Italy] Model 2350 with a Chrompack micro-thermoconductivity detector, described by Börjesson and Svensson, 1997a).
Soil Incubations
In order to determine 13C–CH4 discrimination by the methane-oxidizing bacteria (, defined below), and the consumption of CH4 and production of CO2, incubations of cover soil samples were performed. In these incubations we determined both concentration and the 13C of CH4 and CO2 as a function of time. The simultaneous determination of concentration and isotopic composition allowed us to determine the quantities of CO2 produced from methane oxidation and soil organic matter, respectively. Soil was taken from different depths in the profiles, where optimal methane oxidation activity could be expected (Table 2). Following collection the samples were refrigerated at 4°C. Within 24 h, samples were sieved (4-mm mesh), and aliquots of 100 g soil at field moisture water contents were put in 1.1-L Simax glass jars (Sklarny Kavalier, Sazava, Czech Republic), which were sealed with screw caps. Additional ambient air (100 mL) was added to each jar to check for leakage and allow sample removal. Methane (50 mL; EC 200 812-7, Air Liquide, Kungsängen, Sweden) was added at time zero, thus giving an initial gas mixture of 3.9% CH4 and 20% O2. Three gas samples of 0.3 mL were withdrawn and immediately analyzed on a gas chromatograph with a flame-ionization detector (Packard-Becker [Delft, the Netherlands] 428; Börjesson and Svensson, 1997a). Next, 10-mL gas samples were taken and stored on pre-evacuated glass vials for later isotopic and concentration analysis. By removing additional samples, CH4 consumption was followed until at least 80% of the CH4 was consumed. At least eight measurements were done for each incubation study.
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Table 2. Water content in the samples from the Hökhuvud and Falköping landfill cover soils that were used for incubations (mean values for n = 3)
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Soils sampled at Hökhuvud and Falköping 19 and 20 Aug. 1997 were incubated at 25°C, soils sampled at Falköping 24 Feb. and at Hökhuvud 6 Mar. 1998 were incubated at 4°C. The water content of the soil samples was measured gravimetrically (Table 2).
Analysis of Carbon Isotopes
Samples of chamber atmosphere, landfill gas, and incubation samples were analyzed for 13C by injecting 0.1 to 0.5 mL of sample into a gas chromatograph interfaced to a Finnigan (Bremen, Germany) MAT Delta S isotope ratio mass spectrometer inlet system via a combustion interface (Merritt et al., 1995). Chamber air samples were corrected for admixture with background air methane by mass balance. The gas chromatograph was equipped with a Poraplot Q column (Hewlett–Packard, Palo Alto, CA) with a column head pressure of 10 psi (70 kPa). The Finnigan MAT combustion column was operated at 960°C.
Isotopic values are given as 13C () according to the formula:
 | [1] |
where R is the ratio between 13C and 12C isotopes in samples and standard (Peedee Belemnite), respectively.
Soil samples were sent to the Isotopic Service Laboratory (Los Alamos, NM) where they were analyzed for 13C by combustion on an elemental analyzer coupled to a VG-Isomass isotope ratio mass spectrometer (Micromass, Manchester, UK). Reproducibility was 0.1 for 13C.
Calculations of Fractionation Factors and Methane Oxidation
From the incubations of soil with CH4 (described above), 13C values were compared in order to determine the fractionation factor . The term is defined as the ratio of the rate constants of CH4 oxidation, assuming first order kinetics:
 | [2] |
where kL and kH refer to the rate constants of the light (12CH4) and heavy (13CH4) isotopes.
A time series was performed to determine the rate of methane consumption and the change in the methane isotopic signature in the flasks. Fractionation factors () were calculated following the approach derived in Coleman et al. (1981) using their Eq. [14]:
 | [3] |
where M/Mo is the fraction of methane remaining at time t, 13Ct is the 13C value of the methane remaining at time t, and 13Ct=0 is the 13C value of the methane at the initial time. When 13C is plotted versus ln(M/Mo), the slope of the line fit to the data is 1000(1/ - 1).
The proportion of CH4 oxidized as it escaped the landfill through the soil, fox, was calculated as described by Liptay et al. (1998). Generally we will express this fraction as a percentage, (fox x 100). This factor represents the portion of the CH4 flux from the anaerobic zone to the oxic soil layer, which is oxidized:
 | [4] |
where E is the isotope ratio in emitted CH4, A is the isotope ratio in anaerobic-zone CH4, ox is the fractionation factor calculated from incubations (Eq. [3]), and trans is the fractionation occurring during diffusive transport through the cover. For landfills, trans = 1 is assumed, since gas transport through the soil cap is dominated by advection due to pressure differentials (Liptay et al., 1998; Bergamaschi et al., 1998). This assumption means that the 13C value of methane within the anoxic zone is what enters the oxidation zone. Chanton and Whiting (1996) have demonstrated that in advective transport, isotopic fractionation is minimized. We further assume that the 13C value of residual methane due to oxidative microbial processes within the soil is what is emitted to the atmosphere and captured in our chambers. This assumption is reasonable, as Bergamaschi et al. (1998) observed that the 13C value of CH4 captured in surface accumulating chambers was similar to the samples collected closest to the soil–air interface. Bergamaschi's observations are additional supporting evidence that trans = 1.
In soil incubations the CO2 produced from CH4 oxidation was calculated from the total CO2 production, which also included respiration of bulk soil organic matter. We did this by determining the isotopic composition of methane, which had been oxidized as shown below:
 | [5] |
where 13CCH4-ox represents the isotopic composition of methane that is oxidized, 13Cfinal and 13Cinitial represent the final and initial values for the 13C isotopic composition of CH4, and [CH4final] and [CH4initial] represent the final and initial concentrations of CH4. The denominator in Eq. [5] represents the quantity of CH4 oxidized.
The term 13CCH4-ox also represents the 13C of CO2 produced from methane oxidation. However the quantity of methane oxidized (denominator in Eq. [5]) does not represent the amount of CO2 produced from CH4 oxidation, as that carbon can go either to CO2 or into methanotrophic biomass. This split is controlled by the growth efficiency of the bacteria. At high growth efficiency more of the oxidized CH4 will go into microbial biomass. At low growth efficiency more CO2 will be produced. The ultimate fate of oxidized CH4 is important for assessing the contribution of different gases to the total amount of greenhouse gas emissions from landfills to the atmosphere.
The amount of CO2 produced from methane oxidation (CO2CH4ox) can be separated from the CO2 produced from soil organic matter oxidation (CO2SOM-ox) by solving this equation:
 | [6] |
where CO2total and 13Ctotal represent the total CO2 concentration and its isotopic composition and 13CCH4-ox and 13CSOM-ox represent the isotopic composition of CO2 produced from the oxidation of methane and soil organic matter, respectively.
The total amount (CO2total) and isotopic composition of CO2 (13Ctotal) at the end of the experiment were corrected for CO2 that was in the flask at the time 0 by a similar mass balance equation. We also assumed in these calculations that the 13C of CO2 produced from soil organic matter decomposition was identical to the 13C of bulk soil organic matter.
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RESULTS
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Methane Emissions
Methane emissions from the two landfills were of the same magnitude, ranging from -0.92 mg to 11.3 g CH4 m-2 h-1 at Hökhuvud and from -0.98 mg to 29.6 g CH4 m-2 h-1 at the new landfill in Falköping (Table 3). The low CH4 emissions measured in Falköping on 24 February were probably due to a higher intensity of gas extraction and the fact that the surface was wetter at this time, which directed the CH4 emissions to a few hot spots not covered by the randomly distributed chambers. Therefore, the additional sampling in March was made only on spots with elevated CH4 concentrations as located by Gas-trac scanning.
In summer, the slopes of the landfills had higher CH4 emission rates than the tops, both at Hökhuvud (P = 0.096) and at Falköping (P = 0.0003). In winter, the situation at Hökhuvud was the opposite: there were fairly high emission rates at the top of the landfill compared with the slope. At Falköping, none of the eight chambers on the top showed significant CH4 emissions, while only one chamber out of eight standing in the slope had a significant CH4 emission in February (2.47 mg CH4 m-2 h-1).
Climatic Conditions, Soil Moisture, and Soil Organic Matter 13C
Temperatures measured during the sampling dates are given in Table 5. The first sampling in late August 1997 was done during a relatively warm period. Weather was mild also in January, but cold in March (as normal).
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Table 5. Mean temperatures during sampling, 13C in emitted methane, and calculated methane oxidation (± standard deviation). Numbers in parentheses = number of replicates
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The surface soils (0–5 cm) were very dry in summer and therefore no incubations were made with these samples. Further down in the profiles, soil moisture was almost the same over the year (Table 2). Soil organic matter 13C was -20.2 ± 0.4 (analytical error), n = 1 at Hökhuvud (40–50 cm, collected 23 Aug. 1997) and -21.0 ± 1.2 (replicate error), n = 3 at Falköping. This mean represents samples from three depths: 0 to 5 cm, -19.76; 5 to 15 cm, -21.28; and 15 to 30 cm, -22.08.
Anoxic Zone Methane
Gas samples collected from the gas recovery system in the Falköping landfill had 13C–CH4 values of -54.3 on 20 Aug. 1997 and -53.6 on 3 Mar. 1998 (Table 4). There was no significant difference between these dates (P = 0.50), consistent with a lack of temperature variation in the anoxic methane production zone.
Gas samples from auger holes on the old Hökhuvud landfill contained CH4 with 13C values of -43.4 at 120 cm depth on 19 Aug. 1997, and as high as -18.6 at 85 cm depth. These values could therefore not be regarded as representative for the anaerobic zone, because CH4 in these samples had obviously been exposed to oxidation. The values fell far outside the zone of anoxic-zone methane as compiled by Chanton et al. (1999). Samples collected at 95 to 100 cm depth on 26 Jan. 1998 contained -58.6 13C in CH4 (Table 4), while samples from 80 cm depth still contained as high as -49.2 13C–CH4. A value of -58.6 was therefore used as representative of anoxic-zone CH4 from this landfill, assuming no seasonal variation. No seasonal variation was observed at Falköping or at a landfill in Florida, USA (Chanton and Liptay, 2000), as expected due to the constant temperatures within the landfill maintained by the biological degradation processes.
Incubations—Carbon Dioxide to Methane Ratios and Fractionation Factors
Summer CH4 oxidation rates, in soil samples incubated at 25°C, were higher in the Hökhuvud soil (0.76 µol CH4 g dry wt.-1 h-1; standard deviation = 0.048, n = 4) than in the Falköping soil at both the 15- to 30-cm (0.28 µol CH4 g dry wt.-1 h-1; s.d. = 0.024, n = 4) and at the 5- to 15-cm depth level (0.44 µol CH4 g dry wt.-1 h-1; s.d. = 0.11, n = 2; see, for contrast, average values in Table 6). Examples of incubation data are shown in Fig. 1
. The winter samples, incubated at 4°C, showed significantly lower rates of CH4 oxidation compared with summer. Methane 13C became increasingly 13C enriched in these closed-system incubations as 12CH4 was consumed at a slightly faster rate than 13CH4 by the methane-oxidizing bacteria. The rate of CO2 increase was observed to be considerably less than the rate of methane decrease in these incubations. The isotopic composition of CO2 was intermediate in value between the isotopic composition of consumed methane, calculated by Eq. [5], and the measured isotopic composition of soil organic matter, both of which supported microbial growth and respiration and/or CO2 production. From the concentration and isotopic data, we calculated the quantity of CO2 produced from methane oxidation using Eq. [6]. The ratio of CO2 produced per amount of CH4 consumed varied from 0.15 to 0.37. Apparently, most of the consumed CH4 went into microbial biomass, indicating a high growth efficiency for these bacteria. This parameter is important for assessing the contribution of different gases to the total amount of greenhouse gas emissions from landfills to the atmosphere.
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Table 6. Methane consumption rates, carbon utilization, and values in soil samples incubated at two different temperatures with 3.9% CH4 (± standard deviation)
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The values were calculated from the slopes of the lines in Fig. 2
(using Eq. [3]). Values were higher for the Hökhuvud soil compared with the Falköping cover soil (P = 0.0067). Also, the values were higher for the 4°C incubations than those for the 25°C incubations for both soils (P = 0.0002 at Hökhuvud, P = 0.076 at Falköping, but P = 0.0098 with one outlier excluded; Table 6). If we assume a linear relationship between and temperature, as was reported by Chanton and Liptay (2000), the estimated coefficients for the sloping lines, which represent the change in per degree of temperature, were -0.000448 K-1 for Hökhuvud and -0.000169 K-1 for the Falköping soil.
13C in Emission Chambers
The 13C of CH4 emitted into chambers ranged from -24.0 to -54.8 at Hökhuvud and from -33.5 to -50.9 at Falköping in August 1997. In January 1998 values ranged from -37.5 to -57.9 at Hökhuvud and in March 1998 chamber CH4 had 13C within the ranges -56.1 to -65.8 at Hökhuvud and -53.4 to -59.8 at Falköping. Mean values are given in Table 5.
Methane oxidation was calculated from each of the individual chambers and averaged (Table 5). The chambers with the highest CH4 emission rates had corresponding 13C–CH4 values in the middle of the 13C–CH4 range for the respective site's chambers (i.e., there was no correlation between oxidation and flux rates).
In summer, the oxidation percentage was high for both landfills, and almost 100% at the top of the old landfill, Hökhuvud. In winter, when temperatures fell below 0°C, CH4 oxidation could not be detected. The 13C in emitted CH4 was not significantly different from the 13C of CH4 collected from anaerobic zones (P = 0.15 at Falköping and P = 0.31 at Hökhuvud; see, for contrast, Table 5).
The differences between seasons were significant for both sites in analysis of variance (ANOVA) (August vs. March; P < 0.0001). There was no correlation between CH4 fluxes and 13C in emitted CH4. Both the highest positive fluxes and the negative fluxes, showing net consumption of CH4, were in the middle of the 13CH4 ranges. There were not any differences between slopes and tops (Table 5).
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DISCUSSION
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The 13C values of CH4 from the anaerobic zone of landfills have been reported to fall within a range of -50 to -61, as reviewed by Bergamaschi et al. (1998) and Chanton et al. (1999). Values measured at the landfills in our study, Hökhuvud (-58.6) and Falköping (-54), lie within or close to this range. We observed no seasonal variation at the Falköping landfill site, confirming the observations of Chanton and Liptay (2000).
Methane emissions from the slope of the old Hökhuvud landfill ceased in winter, similar to what was observed in January 1994 in a previous study (Börjesson et al., 1997a). In a Belgian landfill studied by Boeckx et al. (1996), noncovered spots showed lower emissions in autumn than in summer, apparently because these spots were filled with water in autumn. Similarly, the sloping parts of the old landfill in our study could have either become waterlogged, or water (with subsequent ice formation) could have cut this part off from the rest of the landfill, thus directing fluxes through the top. On the top of the landfill, some 15-yr-old Scots pine (Pinus sylvestris L.) trees could have helped to keep pores open for gas exchange. Another explanation could be that the methane production in the slope, located in a peripheral part of the landfill, was suppressed by low temperatures.
We observed that a large portion of the CH4 consumed during oxidation did not yield CO2. Gommers et al. (1988) calculated the fraction of substrate carbon converted to CO2 during assimilation of organic compounds, based on yield data. For methane assimilated via the ribulose–monophosphate pathway of formaldehyde fixation, the portion of carbon converted to CO2 would be 0.12. In our incubated samples (Table 6), we observed ratios varying from 0.15, close to the theoretical 0.12, to as high as 0.37. These ratios are extremely close to those reported by Börjesson et al. (1998a) using an incubation technique that monitored CO2 production and O2 and CH4 consumption. Four other studies reviewed by Börjesson et al. (1998a) have reported CO2 to CH4 ratios that range from 0.16 to 0.4, so these findings appear quite typical. Apparently, when methane is consumed by methanotrophic bacteria in a landfill, the bulk of the consumed methane goes into bacterial biomass, not additional greenhouse gases. One molecule of CH4 is not substituted for a molecule of CO2 at short timescales, as the bacteria have an extremely high growth efficiency. However, given that our results indicate that the bacteria have such high growth efficiencies, it is somewhat surprising that the 13C of soil organic matter was not more 13C depleted, reflecting the uptake of methane into organic tissue. It may be that the bulk of the population of methanotrophs occurs at depths slightly deeper than where we sampled. Alternatively, the microbial biomass may turnover at a high rate and not be preserved. If this is the case, then the ultimate fate of methane that is oxidized must be CO2.
In both soils, the values for fractionation were significantly higher at low temperatures (Table 6). A higher degree of discrimination of heavy isotopes at low temperatures was also observed for CH4 oxidation in landfill cover soils by Chanton and Liptay (2000) and in forest soils by Tyler et al. (1994). In contrast, Coleman et al. (1981) and King et al. (1989) found the opposite in cultures and tundra soil. The higher values in the samples from the old landfill (Hökhuvud), where varied by -0.000448 K-1 for the temperature dependence (derived from data in Table 6), correspond to samples from a landfill studied by Chanton and Liptay (2000), where varied by -0.000438 K-1 for mulch and -0.000433 K-1 for clay. Similar to our results, Tyler et al. (1994) found linear correlation between temperature and the fractionation factor in forest samples to be between 0.00043 and 0.00049 K-1. The lower values obtained in the samples from the new landfill (Falköping), with a regression coefficient for temperature of -0.000169 K-1 indicate a difference in conditions between the two landfills.
Jahnke et al. (1999) proposed that the type of enzyme responsible for the methane oxidation (methane monooxygenase, MMO) may control the degree of isotopic fractionation, where the particulate MMO, which is membrane bound (with Cu as cofactor), has much higher fractionation than the soluble MMO, which is free in the cytoplasm (with Fe as cofactor). Type I-methanotrophs only express the particulate form, whereas Type II-methanotrophs express both forms. In our experiment, there might have been a difference in the microbial community structure between the new and old landfill soils. Since the soil composition did not make any difference in the study by Chanton and Liptay (2000), one hypothesis resulting from our experiment is that the age of the landfill might influence the composition of the methanotrophic population in the cover soil.
The occurrence of CH4 oxidation varies seasonally, being lower or absent in winter and greater in summer. The absence of CH4 oxidation activity in landfills at air temperatures around 0°C was obvious at both sites. This means that the oxidation takes place only in the surface soils, which are temperature sensitive and not deep down in the profile. The lack of CH4 oxidation at low temperatures has implications both for budgets of CH4 turnover in landfills, which on a global scale should include a latitudinal factor, and for the possibility of using enhanced coverage as a sole management practice to mitigate CH4 emissions from landfills in cold climate zones.
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CONCLUSIONS
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Methane oxidation in landfill cover soils is greater in warmer climates. Climatic variables such as moisture and temperature have been shown to control oxidation (Whalen et al., 1990; Bogner and Spokas, 1993; Börjesson and Svensson, 1997a). Our results are consistent with these studies and suggest that there must be systematic variations in the importance of landfill methane oxidation as a function of season and latitude. Methane oxidation has been shown to be the primary variable controlling methane emissions from landfills where gas collection systems are not in place. This results in the antithetical situation of methane emissions being greater in colder winter temperatures than in warmer summer temperature in contrast to most other methane emission sources, which are greater under warmer conditions. The locus of CH4 emissions can vary seasonally. The degree of isotopic fractionation during methane oxidation is temperature dependent and increases at colder temperatures. Methane-oxidizing bacteria have high growth efficiencies and the majority of methane that is oxidized in a landfill goes initially into cellular biomass.
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ACKNOWLEDGMENTS
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This study was supported by the Swedish Environmental Protection Agency (Contract AFR 202/96), the Swedish National Energy Administration (Contract P10856-1), and by the Southeastern Regional Center of the National Center for Global Environmental Change within the U.S. Department of Energy under Cooperative Agreement no. DE-FC03-90 ER61010. We thank the staff and operators of the landfills for their cooperation in this study. We thank Candace Schwartz, Karen Liptay, and Joanne Lombardi for assistance in the laboratory. We are also grateful to the anonymous reviewers who improved the manuscript.
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[Abstract]
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ABSTRACT
INTRODUCTION
