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Journal of Environmental Quality 30:24-30 (2001)
© 2001 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America

TECHNICAL REPORT
BIOREMEDIATION AND BIODEGRADATION

Potential Mineralization of Four Herbicides in a Ground Water–Fed Wetland Area

Lise Larsen, Claus Jørgensen and Jens Aamand

DHI Water & Environment, Agern Allé 11, DK-2970 Hørsholm, Denmark

Corresponding author (jeaa{at}geus.dk)

Received for publication January 3, 2000.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
Herbicides may leach from agricultural fields into ground water feeding adjacent wetlands. However, only little is known of the fate of herbicides in wetland areas. The purpose of the study was to examine the potential of a riparian fen to mineralize herbicides that could leach from an adjacent catchment area. Slurries were prepared from sediment and ground water collected from different parts of a wetland representing different redox conditions. The slurries were amended with O2, NO-3, SO2-4, and CO2, or CO2 alone as electron acceptors to simulate the in situ conditions and their ability to mineralize the herbicides mecoprop, metsulfuron-methyl, isoproturon and atrazine. In addition, the abundance of bacteria able to utilize O2, NO-3, SO2-4+ CO2, and CO2 as electron acceptors was investigated along with the O2–reducing and methanogenic potential of the sediment. The recalcitrance to bacterial degradation depended on both the type of herbicide and the redox conditions pertaining. Mecoprop was the most readily degraded herbicide, with 36% of [ring-U-14C]mecoprop being mineralized to 14CO2 under aerobic conditions after 473 d. In comparison, approximately 29% of [phenyl-U-14C]metsulfuron-methyl and 16% of [ring-U-14C]isoproturon mineralized in aerobic slurries during the same period. Surprisingly, 8 to 13% of mecoprop also mineralized under anaerobic conditions. Neither metsulfuron-methyl nor isoproturon were mineralized under anaerobic conditions and atrazine was not mineralized under any of the redox conditions examined. The present study is the first to report mineralization of mecoprop in ground water in a wetland area, and the first to report mineralization of a phenoxyalcanoic acid herbicide under both aerobic and anaerobic conditions.


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
GROUND WATER–FED WETLANDS in agricultural areas are potentially exposed to herbicides leaching from sprayed fields. As wetlands often filter ground water from large catchment areas, their inherent ability to retain and degrade herbicides is of vital importance to the recipient streams.

Though partial degradation of herbicides or adsorption to sediment may decrease the measurable concentration of the compounds, the only way to eliminate the xenobiotic compounds from the environment is by complete degradation (mineralization) to CO2, water, and inorganic components. Although some steps in the degradation pathway may be abiotic, mineralization is thought to primarily depend on biotic processes (Alexander, 1981; Heider and Fuchs, 1997). Degradation of a herbicide thus depends on its chemical structure and the geochemical and ecological conditions pertaining in the environment. The chemical structure of herbicides often contains aromatic ring structures and substituents such as halogens, which induce recalcitrance to degradation (Knackmuss, 1996; Gibson, 1968).

Mineralization of herbicides is mostly studied aerobically, though the compounds are often exposed to anaerobic conditions in the environment, particularly under waterlogged conditions such as found in wetlands. Many herbicides are recalcitrant in anaerobic aquifers (Larsen and Aamand, 2000) and knowledge of potential degradation in wetlands is required to assess the risk of contamination.

The degradation pathway partly depends on the redox conditions. For example, while aromatic rings are cleaved by oxidation under aerobic conditions (Smith, 1994; Moorman, 1994), under anaerobic conditions the ring must be reduced to achieve cleavage, thus involving completely different enzymatic processes (Heider and Fuchs, 1997; Schink et al., 1992; Moorman, 1994). Similarly, the dehalogenation of herbicides depends on the redox conditions, with a reducing environment favoring reductive dehalogenation (Fetzner and Lingens, 1994).

The aim of the present study was to investigate the potential of a ground water–fed riparian fen to mineralize herbicides representing four widely used groups, namely the phenoxyalcanoic acid, sulfonylurea, phenylurea, and triazine herbicides. Those studied were mecoprop ((±)-2-(4-chloro-2-methylphenoxy)propanoic acid), metsulfuron-methyl (2-[[[[(4-methoxy-6-methyl-1,3,5-triazine-2-yl)amino]carbonyl]amino]sulfonyl]benzoate),isoproturon (3-(4-isopropylphenyl)-1,1-dimethyurea), and atrazine (6-chloro-N2-ethyl-N4-isopropyl-1,3,5-triazine-2,4-diamine).

Previous studies of the wetland (Blicher-Mathiesen and Hoffmann, 1999; Dahl, 1995) showed that its geochemistry varied such that the ground water was successively exposed to aerobic, denitrifying, sulfate-reducing, and methanogenic conditions. Mineralization of radiolabeled herbicides was thus studied in sediment–ground water slurries amended with electron acceptors to simulate the redox conditions prevailing at the sample site. In order to relate mineralization of the herbicides to the abundance and type of the indigenous bacterial populations and the redox conditions pertaining along the ground water flow path through the fen, we also determined the size of the bacterial populations utilizing each of the available electron acceptors and the turnover of some terminal electron acceptors.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 REFERENCES
 
Site Description and Sampling
The study site was a 3136-m2 freshwater wetland (Fig. 1) near Voldby stream in the river Gjern catchment area in eastern Jutland, Denmark. The vegetation is characterized by a dense layer of herbaceous plants dominated by reed grass [Glyceria maxima (Hartm.) Holmb.] interspersed with alder [Alnus glutinosa (L.) Gaertn.] and willow (Salix spp.). The area is permanently under 10 cm of water, with more than 90% of the water entering as ground water and the remainder as precipitation. The fen is recharged at the bordering hillslope, with a constant influx of ground water from a 114-km2 predominantly agricultural (74%) catchment area. The recharge is lateral and aerobic at the hillslope with the water gradually becoming increasingly reduced as it flows towards Voldby stream (Blicher-Mathiesen and Hoffmann, 1999; Dahl, 1995; Blicher-Mathiesen et al., 1998).



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  		Fig. 1. Topographic map of the wetland area indicating the transect. Ground water is recharged at the hillslope from where it flows parallel to the transect towards the recipient stream, Voldby stream. Samples were taken along the transect at Sites 0 and 5, and at two depths at Site 9. The arrow shows the ground water flow direction (redrawn from Dahl, 1995)

 
Sediment and ground water were sampled at four sites designated S0, S5, S9-I, and S9-II (Fig. 1), with S9-I being located 1.1 m below the surface and S9-II 60 cm below the surface at the same location. These four sites were selected because previous investigations had shown that the dominant electron acceptor at each of the sites differed as follows: S0: O2, S5: NO-3, S9-I: SO2-4, and S9-II: CO2 (Dahl, 1995). The sediment type was sand at Stations 0 and 5, sandy peat at S9-I, and peat at S9-II. Sediment cores were collected using a stainless steel borer and the sediment immediately transferred to sterile air-tight containers, completely filling them. Ground water was collected from polyethylene piezometers nested at the stations using an immersion pump and silicone tubing. Samples were stored at 4°C until use.

Analysis of Ground Water
Oxygen, pH, and conductivity were measured in the field in a flow cell mounted with electrodes: Oxi 320 for oxygen, pH 320 for pH, and LF320 for conductivity measurements (WTW, Weilheim, Germany). Phosphate was measured by a modified ascorbic acid method (American Public Health Association, 1995, p. 4113–4114), nitrite was measured spectrophotometrically using a method modified from Crosby (1967), and other anions were analyzed on a Dionex (Sunnyvale, CA) 2000i ion chromatograph. Ammonium was measured as indophenol (Harwood and Kühn, 1970). Ferrous iron was measured spectrophotometrically (A540 nm) on 10-mL samples, which were fixed immediately with 1.4 mL solution of 2,2-bipyridine (2.50 g L-1) and CH3COONa (580 g L-1) at pH 5.6. For measurement of methane, 6 mL water was sampled in ground water–flushed syringes, injected into Venoject tubes (Terumo, Leuven, Belgium), and analyzed on a Shimadzu (Kyoto, Japan) GC-9A gas chromatograph equipped with a flame ionization detector (FID). Separation of the gas constituents was performed on a 4 m x 3.175 mm (0.125 in) stainless steel column packed with silica gel 70/80 mesh using helium as carrier gas. Alkalinity was determined by titration of 10-mL samples with HCl to pH 4.5. The total organic carbon content of the sediment was determined by combustion (LECO [St. Joseph, MI] CS-200).

Enumeration of Bacteria
The abundance of aerobic, denitrifying, sulfate-reducing, and methanogenic bacteria in the sediment at each site was determined. Aerobic or anaerobic suspensions (according to the conditions at the sampling site) were prepared in a phosphate-buffered saline solution (8.0 g L-1 NaCl, 0.34 g L-1 KH2PO4, 1.21 g L-1 K2HPO4, pH 7.0) and shaken vigorously prior to inoculation. Dilutions were plated on noble agar (Difco, Detroit, MI) containing 300 mg L-1 of Tryptic Soy Broth (Difco) and 25 mg L-1 of pimaricin (Merck, Darmstadt, Germany) and incubated at 21°C for 9 d for aerobic colony counts.

Anaerobic bacteria were enumerated using the three tube Most Probable Number (MPN) method. Denitrifying bacteria were enumerated by a method described by Tiedje (1982) with denitrification being verified both by reduction of nitrate and gas production, and by determination of nitrous oxide after addition of acetylene.

Enumeration of sulfate-reducing and methanogenic bacteria was performed in 25-mL butyl rubber stoppered serum bottles containing 10 mL anaerobic medium (Angelidaki et al., 1990) with the following modifications. Cysteine-HCl was omitted and the medium was prepared under flushing with a mixture of 30% CO2 and 70% N2. Yeast extract (0.25 g L-1, Difco), tryptone (0.25 g L-1, Difco), and sodium acetate (0.28 g L-1) were added as electron donors. The medium for the sulfate-reducing bacteria was amended with 0.4 g L-1 Ca SO4·2H2O, 0.8 g L-1 MgSO4·7H2O, and 0.5 g L-1 FeSO4·7H2O as the electron acceptor (Ludvigsen et al., 1995). The pH was adjusted to 7.0 and the flasks were autoclaved for 20 min at 121°C. The medium was reduced with Na2S·7–9H2O (0.05 g L-1) and inoculated with 10% by volume of the dilution series prepared from the suspensions. For enumeration of methanogenic bacteria, 10 mL hydrogen was added to the bottle headspace with a syringe to serve as an additional electron donor.

Respiration Rates
Aerobic respiration was determined as oxygen consumption in 116-mL serum bottles sealed with butyl rubber stoppers. Approximately 10 g of wet sediment was added to five replicate bottles and incubated at 10°C. A test tube containing 2 mL 1 M NaOH was inserted in each bottle to trap evolved CO2. The pressure in the bottles was measured daily for 1 wk using a pressure transducer (Inst. Grasslands & Environmental Research, Aberystwyth, UK) with a digital readout voltmeter. The pressure transducer had a range of 0 to 15 psi (0–103 kPa) with an accuracy of 0.1 psi ± 2% (Theodorou et al., 1994). Five empty bottles served as controls to compensate for variation in atmospheric pressure. Standard curves were produced by withdrawal of known amounts of air from the bottles.

Potential methanogenesis was determined in similar bottles amended with sediment corresponding to 15 g dry weight and 30 mL ground water. Hydrogen (12.5 mL) and Na-acetate (3.4 mM) were added as electron donors. The flasks were incubated at 10°C and evolution of methane followed on a gas chromatograph for 45 d.

Mineralization Studies
For studies of herbicide mineralization, 100 µL of a stock solution containing the radiolabeled and unlabeled herbicide (Fig. 2) in methanol was added to sterile flasks. This corresponded to a total of 0.75 µg herbicide per flask or ~25 µg L-1 ground water. The flasks contained 1667 Bq of either [ring-U-14C]mecoprop (Amersham Pharmacia [Oakville, ON, Canada], radiochemical purity >98.4%), [ring-U-14C]isoproturon (Amersham Pharmacia, radiochemical purity >97%), or [ring-U-14C]atrazine, (Sigma [St. Louis, MO], radiochemical purity 95%), or 750 Bq [phenyl-U-14C]metsulfuron-methyl (donated by Dupont [Wilmington, DE], radiochemical purity 99.0%). The methanol was allowed to evaporate before addition of the sediment. Flasks (100 mL) with ground glass joints were used for the aerobic experiments, while 116-mL butyl rubber stoppered serum bottles with aluminum crimp seals were used for the anaerobic experiments.



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  		Fig. 2. Chemical structure of (A) mecoprop, (B) metsulfuron-methyl, (C) isoproturon, and (D) atrazine

 
The slurries were prepared in triplicate using wet sediment corresponding to 10 g dry weight and 30 mL ground water sampled from the same depth as the sediment. The sediment–ground water slurries from site S0 were set up aerobically while those from site S5 were set up in an atmosphere of N2 with the addition of 10 mM KNO3. The slurries from sites S9-II and S9-I were set up anaerobically in an atmosphere of 30% CO2 and 70% N2. The flasks simulating sulfate-reducing conditions were amended with 0.3 mL stock solution of Na2SO4 to supply 10 mM SO2-4 in the slurries (not taking into account the water added with the wet sediment). Slurries simulating methanogenic conditions were not amended with additional electron acceptors.

To trap the 14CO2 generated by mineralization of the radiolabelled herbicide, a test tube containing 1 mL 1 M NaOH was mounted vertically in the culture flasks held in place by a metal needle inserted into the butyl rubber stopper. The NaOH was changed at each sampling occasion. The NaOH solution in the aerobic flasks was sampled using a pipette in a flow chamber while that in the anaerobic flasks was sampled anaerobically through the stopper using a N2–flushed disposable syringe mounted with a valve and a 0.9- x 120-mm stainless steel needle (Unimed, Lausanne, Switzerland). We added 10 mL OptiPhase HiSafe scintillation fluid (Wallac, Loughborough Leics, UK) to each sample and the radioactivity determined using a Wallac 1409 liquid scintillation counter. Calculations on cumulative mineralization are corrected for quenching and background decay. The data reported here represent the average 14CO2 evolution for triplicate slurries.

Radiolabeled 14CH4 in the headspace at the end of the experiment was combusted and measured as 14CO2. A 5-mL sample of headspace was passed through the flame ionization detector of a gas chromatograph and trapped in ethanolamine and 2-methoxyethanol (1:7 v/v) (Iversen and Blackburn, 1981). To check for 14CO2 in the headspace of the slurries, the samples were also injected while the flame ionization detector was switched off.


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 REFERENCES
 
Geochemistry and Microbiology
Analysis of the water chemistry at the sample sites verified that the ground water became increasingly reduced as it flowed toward the recipient stream and that the sediment and ground water samples in fact represented different redox zones (Table 1). Oxygen at aerobic levels was only detected at site S0, located on the hillslope. Sites S0 and S5 closest to the ground water recharge were rich in nitrate (>1 mM), whereas sites S9-I and S9-II further along the transect were anaerobic and depleted of nitrate. The sulfate concentration was 927 µM at S9-I but only 94 M at S9-II, where methane was observed at a concentration of 24 mM. The presence of ammonium and ferrous iron at S9-I and S9-II further indicated the reduced state of these sites.


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  		Table 1. Geochemistry of the sampling sites, March 1998

 
Bacterial Abundance and Respiration Rates
Aerobic bacterial abundance was one order of magnitude greater in the aerobic sediment (S0: 5.8 x 106 CFU) than in the denitrifying sediment (S5: 6.4 x 105 CFU) (Table 2). The abundance of denitrifying bacteria was lower than that of aerobic bacteria under both aerobic (S0) and denitrifying (S5) conditions. The number of bacteria using the prevalent electron acceptors was lower in the sulfate-reducing (S9-I) and methanogenic (S9-II) sediment (Table 3) than in the aerobic (S0) and denitrifying (S5) sediment. Both sulfate-reducing and methanogenic bacteria were more abundant in the methanogenic sediment (S9-II) than in the sulfate-reducing sediment (S9-I). However, the potential rates of methanogenesis did not differ significantly between these two sediments.


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  		Table 2. Bacterial density and activity at the aerobic and denitrifying sites

 

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  		Table 3. Bacterial density and activity at the sulfate-reducing and methanogenic sites

 
Mineralization of Herbicides
Mecoprop proved to be the most readily mineralized of the four herbicides and the only one to be mineralized under anaerobic conditions (Fig. 3) . Under aerobic conditions (S0 slurries) mineralization started immediately, and 36% of the herbicide had mineralized by the end of the experiment, 473 d later. Under anaerobic conditions (S5 slurries), in contrast, only 8% of the mecoprop mineralized during the course of the experiment. Similarly low rates of mineralization were obtained under both sulfate-reducing (S9-I: 10%) and methanogenic conditions (S9-II; 13%). Moreover, mineralization of mecoprop in both S9-I and S9-II slurries was independent of the presence of added sulfate since the omission of sulfate from S9-I slurries and the addition of sulfate to S9-II slurries did not significantly affect the mineralization of mecoprop (data not shown). Also, there was no significant difference between the amounts of accumulated 14CO2 from mecoprop in anaerobic slurries amended with nitrate and anaerobic slurries with sulfate or without additional electron acceptor. The presence of anaerobic conditions in the S9-I and S9-II slurries was confirmed in two ways. (i) To ensure that oxygen had not entered the slurries during sampling, methane was measured in the culture flasks that had been sampled during the experiment and compared with unsampled controls. There was no significant difference in methane concentration in the sampled and unsampled slurries except for the S9-II slurries containing atrazine (data not shown). Hence, methanogenesis was not inhibited due to oxygen entering the slurries during sampling. (ii) Radiolabeled methane in the headspace of slurries was transformed to 14CO2 and measured. Approximately 0.1 to 1.5% of the 14C initially added as [ring-U-14C]mecoprop was recovered as 14CH4 after 460 d, corresponding to 1.0 to 12.3% of the accumulated 14CO2.



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  		Fig. 3. Mineralization of [ring-U-14C]herbicides to 14CO2 in the presence of various electron acceptors. ({blacksquare}) S0, aerobic; ({diamondsuit}) S5, anaerobic + NO-3; ({blacktriangleup}) S9-I, anaerobic + SO2-4 + CO2; (x) S9-II, anaerobic + CO2

 
Mineralization of metsulfuron-methyl under aerobic conditions (S0 slurries) started after a 10-d lag phase, with mineralization developing exponentially until Day 360 and reaching 29% by Day 473 (Fig. 3).

Mineralization of isoproturon under aerobic conditions was slower than that of the other herbicides (Fig. 3), although it started immediately. By the end of the experiment, only 16% of the herbicide had been mineralized.

The mineralization of atrazine was less than 1.6% under anaerobic conditions and 4.5% in aerobic slurries. That is below or very close to the level of radiochemical impurities in atrazine, 3%. The level of radiochemical impurities sets the detection limit for the mineralization experiments, as mineralization of herbicide is indistinguishable from mineralization of impurities when the concentration of accumulated 14CO2 is within the concentration of radiochemical impurities. Also, the mineralization of metsulfuron-methyl and isoproturon was below the level of radiochemical impurities under the three anaerobic conditions applied. Therefore, it was concluded that atrazine was not mineralized under any of the conditions applied, and metsulfuron-methyl and isoproturon were not mineralized anaerobically.


   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
REFERENCES
 
The similarity in the abundance of aerobic and denitrifying bacteria in the denitrifying sediment (S5) may reflect the facultative nature of denitrifiers (Tiedje, 1988). Thus the bacteria may actually represent the same bacterial population able to grow under both aerobic and denitrifying conditions. In contrast, the higher abundance of aerobic than denitrifying bacteria in the aerobic sediment (S0) probably indicates that most of the bacteria are obligate aerobes.

The greater abundance of sulfate-reducing and methanogenic bacteria in the methanogenic sediment (S9-II) than in the sulfate-reducing sediment (S9-I) may be due to the higher organic matter content of the methanogenic zone, which was 0.50 m above S9-I. While the potential rate of methanogenesis was similar in the S9-I and S9-II sediment slurries under experimental conditions, field measurements revealed a considerably higher methane concentration at Site S9-II than at S9-I. The discrepancy may be due to the fact that there is a continuous input of sulfate at Site S9-I with resultant inhibition of methanogenesis, whereas the sulfate becomes exhausted under the experimental conditions.

While mineralization of mecoprop has previously been observed in unsaturated sediments (Reffstrup et al., 1998; Helweg, 1993), the present study is the first report of mecoprop mineralization in ground water in a wetland area. The findings indicate that mineralization also occurred under anaerobic conditions, something not previously reported for a phenoxyalcanoic acid herbicide. Although the degradation pathway for mecoprop has not yet been fully elucidated, under aerobic conditions it may follow that of the related herbicide 2,4-D (2,4-dichlorophenoxyacetic acid). The latter is dealkylated and hydroxylated to a cathecol, which is ortho-cleaved by the action of a dioxygenase (Tiedje et al., 1969; Pieper et al., 1988). This pathway can obviously only proceed in the presence of oxygen. Although complete degradation of phenoxyalcanoic acids has not been reported under anaerobic conditions, initial transformation (i.e., reductive dechlorination and dealkylation) has been observed for 2,4-D and another phenoxyalcanoic acid, 2,4,5-T (2,4,5-trichlorophenoxyacetic acid)—in the latter case also in aquifer sediment (Gibson and Suflita, 1990; Mikesell and Boyd, 1985). Other monoaromatic compounds are dearomatized by reduction under sulfate-reducing and methanogenic conditions (Heider and Fuchs, 1997), and reduction could be a potential pathway in the mineralization of mecoprop observed in this study.

The mineralization of metsulfuron-methyl followed a logarithmic trend, possibly indicating bacterial growth (Alexander and Scow, 1989). The herbicide may have supported enrichment of bacteria specifically degrading the phenyl ring of metsulfuron-methyl. The enrichment of a small or slowly growing population of degrading bacteria could explain why significant mineralization of metsulfuron-methyl was only observed after 360 d. While the present results indicate mineralization of the phenyl ring of metsulfuron-methyl, the triazine ring may still persist since cleavage of the urea bridge is an early step in the degradation of metsulfuron-methyl, dividing the herbicide into a sulfonamide and a triazine derivative (Berger and Wolfe, 1996).

Isoproturon was mineralized under aerobic conditions, as reported previously for topsoils (Johnson et al., 1998; Pieuchot et al., 1996; Perrin-Ganier et al., 1995; Kubiak et al., 1995). In those studies, the mineralization rate was generally higher, however, with 10 to 39% being mineralized within approximately 2 to 3 mo as compared with only 14% after 1 yr in the present study. The fate of isoproturon under anaerobic conditions does not appear to have been reported previously.

Our finding that ring-labeled atrazine was not mineralized under any of the redox conditions examined is in concert with numerous reports of the inability of sediments to mineralize atrazine (Kruger et al., 1997, 1993; McMahon et al., 1992; Dousset et al., 1997). Aerobic soils are able to mineralize the herbicide, however (Radosevich et al., 1996; Wolf and Martin, 1975; Vanderheyden et al., 1997; Topp et al., 1995), and mineralization of atrazine has been reported in pure cultures under both aerobic and anaerobic conditions (Radosevich et al., 1995; Mandelbaum et al., 1995; Shapir et al., 1998).

The average residence time of the ground water in this wetland area is 40 to 50 d (Dahl, 1995), a short period relative to the herbicide mineralization rates detected. Moreover, as only a small fraction of the wetland area is aerobic, its buffering capacity with regard to herbicides leaching from the catchment area must be negligible. Nevertheless, some herbicide retention may occur by sorption, and some initial transformation of herbicides to metabolites may occur before they enter the recipient stream.


   ACKNOWLEDGMENTS
 
This work was funded by the Danish Environmental Research Programme, SMP 96. We gratefully acknowledge the skilled technical assistance of Spire Kiersgaard (GEUS) and Mette Albrektsen (DHI), as well as the assistance of Lars Kyhnau Hansen, Department of Geology and Geotechnical Engineering, Technical University of Denmark, with the 14C-methane measurements.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 




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