Dichlorodifluoromethane

Biodegradation in soil

Link to relevant study record(s)

Reference

Reference 1

Endpoint:

biodegradation in soil: simulation testing

Type of information:

experimental study

Adequacy of study:

key study

Study period:

Not specified

Reliability:

2 (reliable with restrictions)

Rationale for reliability incl. deficiencies:

other: Taken from publically available data, and is considered accurate based on the registrants experience of the substance. The study is well documented.

Qualifier:

no guideline followed

Principles of method if other than guideline:

See "details on experimental conditions" listed below.

GLP compliance:

not specified

Test type:

other: inherent biodegradability

Radiolabelling:

no

Oxygen conditions:

aerobic/anaerobic

Remarks:

anoxic & oxic conditions

Soil classification:

not specified

Soil no.:

#1

Soil type:

other: Peat

% Org. C:

> 85

pH:

4.5 - 4.8

Bulk density (g/cm³):

0.07

Details on soil characteristics:

Five intact peat cores (10 cm diameter; 30 cm depth) were collected from a conifer swamp (Labrador Hollow) near Truxton, New York, USA, on 29 June 1993 using PVC pipe corers (40 cm length). Briefly, the site has a canopy of white pine (Pinus strobus), red maple (Acer rubrum), larch (Larix laricina) and an open understory of ericaceous shrubs, with a well-developed ground cover of Sphagnum mosses (S. girgensohnii, S. russowii, S. henryense). The peat deposit is at least 3 meters deep and is acidic (pH = 4.8) with a low bulk density (0.07 g cm-3) and high content of organic matter (>8S%). The water table fluctuates from the surface of the peat to about 20 cm below the peat surface during the growing-season. The climate of the region is temperate continental with short, cool summers. Mean annual temperature is 6.0°C with a mean monthly temperature of -5.6°C in January and 20.3°C in July. Mean annual precipitation is 1013 mm, coming as snow and rain.
Three intact peat cores (10 cm diameter; 30 cm depth) were collected from a black spruce-tamarack bog near Rome, New York, USA on 15 June 1994 using PVC pipe corers (40 cm length). This 20 ha bog is located in the Rome Sand Plains, which is a 10,000 ha area of mostly upland forest with an assortment of sand dunes and peat bogs within it. The vegetation of the bog is a mat of Sphagnum rubellum with clumps of black spruce (Picea mariana), L. laricina and pitch pine (Pinus rigida) and ericaceous shrubs; A. rubrum is also prevalent. The peat deposit is at least 4 m deep and is acidic (pH = 4.5) with high organic matter (>85%). The site is relatively dry and no open water is evident. The climate is similar to Labrador Hollow .

Soil No.:

#1

Duration:

10 - 36 d

Soil No.:

#1

Initial conc.:

1 other: ppmv

Based on:

test mat.

Details on experimental conditions:

Chemicals: The gas mixture contained 1.5 ppmv CFC-11, 1.0 ppmv CFC-I2 and 1.0 ppmv SF6. We diluted the mixture for some of the studies. The SF6 was included as an inert tracer to monitor possible gas leakage.
Incubation of Whole Peat Cores: Each intact peat core was held briefly (3 days at 0°C then at 22°C for swamp peat; 5 days at 12°C for bog peat) before placing a plastic cap over each core to seal a headspace of about 1l. The existing headspace was replaced with N2 -- and 50 ml of the SF6-CFC gas mixture was injected into the headspace of the swamp peat cores, and 25 ml of the SF6-CFC gas mixture was injected into the headspace of the bog peat cores. The cores were kept at 22°C, and gas concentrations in the headspace were sampled daily (7 d for the swamp peat cores; 10 d for the bog peat cores) through a septum in the cap using a gas-tight syringe. The sample was replaced with an equal volume of N2 to maintain a constant gas volume/pressure in the headspace.
Incubation of Peat Samples: After completing the whole-core sampling, the peat cores were extruded from their PVC cylinders and several subsamples (90-100 g) from the 0-10 and 20-30 cm depths were sealed in 1l volume Mason jars. The following treatments were established in triplicate samples per core and per depth for peat from both sites: (i) anoxic -- 100 ml of H20 was added, and the headspace was replaced with N2; (ii) oxic -- the peat was exposed to room air for 2 weeks while the initial moisture content was maintained, and the headspace was room air; and (iii) autoclaved -- the peat was autoclaved overnight for 3 successive nights, and the headspace was replaced with N2. Three additional treatments were established in triplicate samples per core and depth for the bog peat subsamples: (iv) methanogen inhibited -- 75 ml of H20 was added along with 2-bromoethyl sulfonic acid (BES - an inhibitor of methanogen activity), and the headspace was N2; (v) carbon added -- 75 ml of H20 was added along with 7 ml of dextrose solution (0.45 g dextrose in 200 ml H20), and the headspace was N2; (vi) CH4 added -- 40 ml of pure CH4 replaced 40 ml of the headspace, the headspace was room air.
Treatment (i) was set up as the anoxic condition most similar to that of the peatlands. Treatment (ii) was set up as the oxic counterpart to (i). In treatment (iii) all the organisms were killed and enzymes destroyed by autoclaving in order to determine if CFC consumption in anoxic conditions was biological. Treatment (iv) was set up to determine the role of methanogens in CFC consumption. Treatment (v) provided an abundant energy source (carbon in the form of dextrose) for anaerobic microorganisms; while treatment (vi) was set up to enhance the activity of methanotrophic bacteria.
In addition, we injected the SF6-CFC mixture into three jars without peat to serve as controls for possible gas leakage from jars. One set of jars including all of the treatments was incubated at 6°C, and second complete set was incubated at 22°C. The headspace was sampled periodically during the incubation period (40 days for the swamp peat; 46 days for the bog peat), and N2 was added after each sample to maintain gas volume/pressure in the jars. Following the last sample, headspace volume was determined and the peat samples were dried to determine the peat mass.

% Degr.:

23

Parameter:

test mat. analysis

Sampling time:

10 d

% Degr.:

85

Parameter:

test mat. analysis

Sampling time:

36 d

Transformation products:

not measured

Evaporation of parent compound:

not measured

Volatile metabolites:

not measured

Residues:

not measured

Details on results:

Incubation of Whole Peat Cores: The concentration of SF6 remained constant in the headspace above the bog peat cores throughout the 10-d sampling period; hence, no leakage occurred from the cores. This was largely true for the swamp peat cores, even though the SF6 concentration fluctuated a bit from one sampling date to the next; overall, there was no significant (P > .05, repeated-measures ANOVA) change in concentration during that time. The concentration of CFC-12 in the headspace above the swamp peat cores also showed no significant (P > .05) change in concentration, while the CFC-12 concentration in the headspace above the bog peat cores decreased slightly (23% after 10 d; P < 10).
Incubation of Peat Samples at 22°C: The concentration of CFC-12 decreased by 85% after 36 d, but only in the jars with surface peat from the conifer swamp incubated under anoxic conditions; otherwise, the CFC-12 concentration remained fairly constant throughout the incubations.
Rates of CO2 production and CH4 production by peat samples were determined during the first two days of the incubation period in order to estimate microbial activity and, in particular, methanogen activity. Peat from the conifer swamp had significantly (P < .01, t-test) higher rates of CO2 production for samples from the 0-10 than from the 20-30 cm depths, whereas the deeper peat had higher rates of CO2 production than surface peat from the bog. Likewise, the rate of CH4 production by swamp peat samples incubated under anoxic conditions was significantly (P < .01) higher for peat samples from the 0-10 than from the 20-30 cm depths, and vice versus for the CH4 production rates among depths for peat samples from the bog. Rate constants were about 10 times smaller at 6°C than at 22°C for CFC consumption in any treatment.

Results with reference substance:

No data.

Table 1. Production rates for CO2and CH4by peat samples incubated under either anoxic or oxic conditions at 22°C. Numbers are means (n = 5 for swamp; n = 3 for bog). Depth CO2(μmol g^-1dry peat h^-1) CH4(nmol g^-1dry peat h^-1) (cm) Anoxic Oxic Anoxic Oxic Swamp 0-10 0.50 0.71 11.5 <0.1 20-30 0.12 0.12 1.8 <0.1 Bog 0-10 0.12 0.33 4.0 <0.1 20-30 0.25 0.43 6.8 <0.1 Table 2. First-order rate constants (k d^-1) for consumption of CFC-12 by intact peat cores and peat samples from those cores incubated under different conditions. Numbers are means (n = 5 for swamp; n = 3 for bog). Incubation Whole core Surface (0-10 cm) Deep (20-30 cm) CFC-12 Anoxic Swamp 0.0552 0.0593 0.0078 Bog 0.0007 0.0013 0.0018 Oxic Swamp N.D. <0.0005 <0.0005 Bog N.D. 0.0053 0.0025 Autoclaved Swamp N.D. <0.0005 <0.0005 Bog N.D. 0.0029 0.0013 CH4Inhibited Bog N.D. 0.0022 0.0041 Carbon added Bog N.D. 0.0020 0.0024 CH4added Bog N.D. 0.0008 0.0012

Conclusions:

Some biodegradation noted under anaerobic conditions. The substance achieved 23% degradation in 10 days and 85% in 36 days.

Executive summary:

The study evaluated the potential consumption of chlorofluorocarbons CFC-11 and CFC-12 by peat soil from a conifer swamp and a temperate bog in New York State in order to assess whether extensive northern peatlands might serve as a sink for atmospheric CFCs. Intact peat cores maintained with an anoxic headspace over the peat surface consumed CFC-11 and minor amounts of CFC-12. The consumption of CFC-11 showed a first order rate constant of 0.122 d-1; hence, molecular diffusion transports CFCs through vegetation to consumption sites in the peat. Peat samples from the 0-10 cm depth in the swamp site showed higher microbial activity and consumed both CFC-11 and CFC-12 at higher rates than deeper peat samples (20-30 cm depth). Conversely, deeper peat from the bog showed higher consumption rates for CFCs than peat samples from the 0-10 cm depth corresponding to the higher microbial activity deep in the peat profile. For both CFCs, consumption by incubated peat samples followed the series: anoxic > anoxic with carbon added >> autoclaved > oxic = anoxic methanogen inhibited > oxic with CH4. The results suggest that anaerobic soil in northern peatlands have the capacity to serve as a sink for atmospheric CFCs. The substance achieved 23% degradation in 10 days and 85% in 36 days respectively.

Reference 2

Endpoint:

biodegradation in soil

Type of information:

experimental study

Adequacy of study:

supporting study

Study period:

Not specified

Reliability:

2 (reliable with restrictions)

Rationale for reliability incl. deficiencies:

other: Taken from publically available data, and is considered accurate based on the registrants experience of the substance.

Qualifier:

no guideline followed

Principles of method if other than guideline:

See "details on experimental conditions" listed below.

GLP compliance:

not specified

Test type:

other: ready biodegradability

Specific details on test material used for the study:

Details on properties of test surrogate or analogue material (migrated information):
Not specified

Radiolabelling:

not specified

Oxygen conditions:

anaerobic

Soil classification:

not specified

Details on soil characteristics:

The digesters were filled with MSW samples (sieved fraction with < 10 cm), biowaste or compost.The MSW was unsorted waste from households, trade and communities, and the organic waste came from private households of Hamburg.

Soil No.:

#1

Duration:

140 d

Details on experimental conditions:

The CFCs R11, R12, and R113 and the VCCs tetrachloroethylene, 1,1,1-trichloroethane, tetrachloromethane, and dichloromethane and their metabolites in the gas phase of the test digesters were examined with standard analytical methods. In order to control the environmental conditions, the gas composition (methane, carbon dioxide, hydrogen, oxygen, nitrogen) was analyzed.
The test digesters made of high-grade steel consisted of a 1m long cylinder with and inside diameter of 0.4 m (volume: 126 litres). The digesters had several valves where samples were taken or substances were added (STEGMANN, 1981). For this purpose the valves were sealed gas tightly with septa, so that the sampling and the addition of substances could be made directly from and into the digester with suitable syringes. A probe was attached to one of the valves which showed several drillings with a diameter of 2 mm, starting 10 cm from the top. The perforated part of the probe penetrated into the 10 cm-high layer of (washed) gravel. Via this probe, the VCCs and CFCs added could be distributed in the gravel layer and could flow evenly through the material on top of it.
The digesters were filled with MSW samples (sieved fraction with < 10 cm), biowaste or compost. The MSW was unsorted waste from households, trade and communities, and the organic waste came from private households of Hamburg (Germany).
Some digesters were filled in addition with compost to accelerate the acidphase. The material was only slightly compressed (vibration density 0.25-0.45 t moist/m3 material vol.) in order to ensure gas permeability and to minimize the wall effect on flow. Water content was set at 50-60% wet mass.
The typical milieu parameters of a 126-liter test digester can be described as follows: In the acidphase the pH-value ranged about 5.5 in the leachate. The concentration of organics is high and the ratio of BOD5/COD (biological oxygen demand in 5 days / chemical oxygen demand) is greater than 0.4. The carbon dioxide concentration amounted to 70-90 vol.-% and the methane concentration is 10-30 vol.-% in the gas phase. In the methanephase the pH-value ranged around 7 and the ratio of BOD5/COD is less than 0.1. The concentrations of volatile organic acids are low. The value of carbon dioxide concentration lies at circa 40 vol.-% and the value of methane concentration at 60 vol.-% in the gas phase. The investigations started each time after the highest gas production.
The chemicals (VCCs/CFCs) were added in a concentration of 5 mg/kg dry mass so that the formation of methane was not inhibited. In setting this maximum VCCs/CFCs concentration, a possible inhibition of the biological processes was to be avoided. The literature quotes the minimum concentration of halogenated methane analogues, where an inhibition of methane production can be seen clearly, at approx. 10 mg/kg dry mass (POLLER, 1990). The same value was assumed for the halogenated ethanes and ethylenes here.
According to their respective gas/water distribution balance, the VCCs/CFCs concentrations in the gas were up to 350 mg/m3 gas.
The substances showed different degradation reactions in the different simulated anaerobic degradation phases that can be observed in municipal solid waste landfills (acid and methanephase).

Under acidic environmental conditions, a degradation of the CFC R11 and the VCCs tetrachloroethylene, 1,1,1- trichloroethane, and tetrachloromethane occurred. R12, R113, and dichloromethane were hardly degraded in the acid phase.
R21, the reductive degradation product of R11, was seen to accumulate. It is assumed that R21 was not decomposed. The metabolites of tetrachloromethane were trichloromethane, dichloromethane and chloromethane. They were measured in the gas phase of a digester filled with biowaste. Dichloromethane, however, was hardly degraded. Also in a digester filled with MSW and compost, dichloromethane could not be degraded significantly during the acid phase.
After the addition of R12, the reductive degradation product R22 could only be detected at low concentrations. This shows that R12 is not degraded substantially under acidic environmental conditions. That applies to R113, too.
Since the concentration of the degradation product chloroethane tended to increase after the addition of 1,1,1- trichloroethane, a dechlorination must be assumed.
Chloroethane as a metabolite of 1,1,1-trichloroethane via 1,1-dichloroethane (which could not be detected due to analytical reasons) was presumably metabolized more slowly or not at all.
The experiments have shown that tetrachloroethylene can be degraded in the acidphase, but the degradation rates are low. The metabolite trichloroethylene was produced and 1,1-dichloroethylene was measured 150 days later, but only in low concentrations .

Evaporation of parent compound:

not measured

Volatile metabolites:

not measured

Residues:

not measured

Details on results:

Biological degradation of VCCs/CFCs in the methanephase: Both R11 and R21 were decomposed. The degradation time of R21 declined each time when R11 was added, so that an adaption of the microorganisms degrading R21 can be supposed.
R113 and R12 was decomposed as well. The metabolite R22 was degraded to only a minor degree or not at all. The degradation products of R113, however, could not be identified for analytical reasons.
Continuous biodegradation of R12 under anaerobic conditions: The continuous gassing of digester 6 filled with compost {late methane phase, material volume 0.084 m3) with 512 mgR12/m3 and a gas flow of 0.5 ml/min (369 μg/d) resulted in a degradation rate of 158 μgR12/m3 material vol./h, at a dwell time of 2,800 hours (117 days). The R22 concentration in the gas phase tended to increase. After the continuous gassing had started, the methane concentration sank by approximately 20 vol.-% due to the dilution and the inhibition of methane formation. In this experiment, the degradation rate was higher for the discontinuous operation (700 μg/m3material vol./h) as well.
Continuous anaerobic biodegradation of R11 and R12: With continuous gassing, the degradation rates achieved were clearly below those achieved with discontinuous gassing, because the analyzed substances R11 (and the decomposition product R21) and R12, in high concentrations, have a toxic effect on the organisms which is responsible for their degradation.
On the one hand, R11 and R21 are more easily degraded than R12 (R22 was probably minimally degraded or not at all degraded anaerobically). On the other hand, they are roughly toxic to the microorganisms to the same extent. The continuous degradation rate of R12 was therefore higher than that of R11 because R22, under anaerobic conditions, is less toxic to microorganisms than R21 .

Conclusions:

The continuous degradation rate of R12 was therefore higher than that of R11 because R22, under anaerobic conditions, is less toxic to microorganisms than R21 .

Executive summary:

The biological degradation of VCCs and CFCs was investigated under simulated conditions of landfills in laboratory test digester. Among these, CFC12 was degraded under anaerobic conditions in addition to the methanogenic bacteria in municipal solid waste and organic wastes.The concentrations, above all those of the fully halogenated volatile hydrocarbons, can decrease under suitable environmental conditions. Firstly, they can be stripped with the LFG that is emitted or extracted from the landfill and, secondly, they can be partly or completely degraded biologically. The reductions rates in the landfill, however, will be lower than those calculated on the basis of the degradation rates in laboratory tests. In particular, this is due to the varying physical conditions in the landfills like temperature and humidity, etc. When mining or reconstructing old landfills today it should thus be considered that there are still VCCs/CFCs which might escape into the atmosphere.

Reference 3

Endpoint:

biodegradation in soil

Type of information:

experimental study

Adequacy of study:

supporting study

Study period:

Not specified

Reliability:

2 (reliable with restrictions)

Rationale for reliability incl. deficiencies:

other: Taken from publically available data, and is considered accurate based on the registrants experience of the substance.

Qualifier:

no guideline followed

Principles of method if other than guideline:

See "details on experimental conditions" listed below.

GLP compliance:

not specified

Test type:

not specified

Specific details on test material used for the study:

Details on properties of test surrogate or analogue material (migrated information):
Not specified

Radiolabelling:

not specified

Oxygen conditions:

anaerobic

Soil classification:

not specified

Details on soil characteristics:

Waste Samples: The potential of anaerobic bacteria to degrade BAs under landfill conditions was tested in presence of three types of waste material: A, B, C.
A: Organic waste collected from Danish households. B: Older pre-disposed waste from an American landfill. C: Waste from a laboratory experimental digester simulating landfill conditions.
The waste materials differed in terms of composition, origin, and age. The methane gas production rate of the different waste types was determined using the method by Hansen et al (Hansen, T. L.; Schmidt, J. E.; Angelidaki, I.; Marca, E.; Jansen, J. l. C.; Mosbæk, H.; Christensen, T. H. Method for determination of methane potentials of solid organic waste. Waste Manage. 2004, 24, 393-400).
The organic waste (waste A) was collected from private Danish households in an area where the citizens as a part of a research program were sorting their waste into different fractions including an organic fraction. The organic waste had the highest methane gas production rate (0.014 g CH4 g-1 waste d-1) of the three types of waste tested, and consisted of well-sorted easily degradable food wastes including fruits, vegetables, bread, rice, corn. The organic material was therefore expected to provide the anaerobic bacteria with a labile carbon source to maximize growth.
Older predisposed waste (waste B) was excavated from an American landfill situated in North Carolina. Although the landfill was producing methane, the gas production rate was lower (0.002 g CH4 g-1 waste d-1) in comparison with the Danish organic refuse, probably due to the higher degree of maturity of the organic material. Unlike waste A, however, waste B may have contained anaerobic bacteria that had been pre exposed to different BAs in the landfill environments and could therefore hold a potential to degrade some of the halocarbons evaluated in these experiments.
Samples from a laboratory experimental digester containing refuse (waste C) was the third type of waste tested. The digester was in its methanogenic phase and contained well-decomposed waste, which was also confirmed by a low methane production rate (0.001 g CH4 g-1 waste d-1).
Microcosms were inoculated with methanogenic digested sludge from either a mesophilic biogas reactor receiving agricultural waste (Hashøj Biogas, Denmark) or a biological sewage treatment plant (Lundtofte sewage treatment plant, Denmark).

Duration:

80 d

Duration:

200 d

Details on experimental conditions:

Degradation of BAs Under Methanogenic Conditions: Anaerobic degradation of selected BAs and associated degradation products was examined in glass microcosm bottles (327mLin total volume) equipped with Teflon-coated butyl rubber septa (8mmthick) held in place by an aluminium screw caps. The septum enabled gas to be sampled or injected by a hypodermic needle and a syringe. A fixed amount of waste (10 g) was placed in each bottle. The waste was homogenized in a blender before use in the experiments. The bottles were inoculated with sludge (40 mL) to ensure anaerobic microbial activity. To obtain anaerobic conditions during microcosm construction, the bottles were flushed with a nitrogen/carbon dioxide mixture (80:20%).
The degradation of halocarbons was studied in single compound tests. Microcosms were spiked with one test compound (CFC-11, CFC-12, HCFC-21, HCFC-22, HCFC-31, HFC-41, HCFC-141b, HFC-134a, or HFC-245fa) using gaseous stock solutions and a glass barrel gastight syringe. Halocarbon initial concentrations, which varied between 150 and 600 μg L-1, were generally selected so they were in the range typically observed in LFG.
The headspace (50 μL) in the microcosms was sampled periodically and analyzed by gas chromatography (GC). Approximately 20 gas samples were extracted during an experimental period (approximately 80-200 days), utilizing less than 1% of the gas volume in the batch microcosm, which was found not to influence the results. Furthermore, a leakage test was performed using argon as a conservative tracer to account for losses through the septum due to perforation during sampling. The test showed that up to 30 gas samples could be extracted over a period of approximately 200 days before the septum started to leak if the syringe was inserted in a new place every time a sample was taken (results not shown). The microcosm experiments were conducted at room temperature (22 °C). All experiments were carried out in triplicate. Sterile control microcosms were prepared to investigate halocarbon loss due to non biological processes (i.e., abiotic degradation, sorption, and volatilization). Control reactors were sterilized by autoclaving (three times for 1 h at 121 °C). Halocarbons were added to the control experiments after autoclaving.
Degradation kinetics were examined by plotting the measured headspace concentrations of the individual compounds versus time and fitting the data to an exponential regression. In all cases, degradation followed first-order kinetics, and degradation rate coefficients were estimated from an exponential fit to the data. For each compound, an average degradation rate coefficient (k1) was calculated based on three replicates. The average degradation rate coefficient (k1), which was based on gas phase measurements, was converted to an aqueous phase rate coefficient (λ) for use as input in the fate model. In this paper, the term degradation is used to describe significant decreases in gas concentration over time in live microcosms relative to sterile controls.
Microbial Mitigation of BA Releases from Foam Insulation Under Methanogenic Conditions: Microcosm experiments with waste mixed with pieces of foam insulation were performed to study whether the presence of anaerobic bacteria can have a mitigating effect on the release of BAs from foam insulation in landfill waste. The experimental procedure was similar to the procedure described for degradation of BAs under methanogenic conditions except that cut pieces of foam insulation were supplemented to bottles instead of a fixed amount of BA. Foam cylinders were cut using a cork bore (1cm diameter and 1 cm height). Four foam pieces cut as cylinders were supplemented to each foam-containing bottle. The experiment also included use of sterile controls to quantify release of BAs in the absence of degradation. Headspace samples were withdrawn and analyzed by GC over time. The experiments ran for 14 weeks at room temperature. Experiments with four BAs (CFC-11, HCFC-141b, HFC-134a, and HFC-245fa) were performed in parallel.

Transformation products:

not measured

Details on results:

Degradation of BAs Under Methanogenic Conditions: Degradation of CFC-11, CFC-12, HCFC-21, HCFC-22, HCFC-31, HCFC-141b was observed in microcosms containing waste inoculated with digested sludge. The degradation was microbially mediated as seen from comparison with the sterilized control batch. In general the degradation followed first-order kinetics. In general, good reproducibility was obtained and results from three replicate batches were almost identical with standard deviations of less than 10% between the obtained degradation rate coefficients.
Methane was produced in all bottles supplemented with waste and inoculum, indicating the presence of active methanogenic bacteria. In experiments supplemented with the lower halogenated HCFCs and HFCs an inhibition of the methane production was observed in comparison with experiments supplemented with CFC-12. No methane production was observed in sterilized bottles with waste and sludge, indicating that the sterilization process was effective.
CFC-12 was degraded in microcosms supplemented with organic household waste. CFC-12 was degraded from ca. 225 μgL-1 to less than 25 μg L-1 within 70 days. The degradation pattern of CFC-12 did not indicate a stoichiometric sequential dechlorination as the sum of the observed degradation products HCFC-22, HFC-32, HFC-41 never exceeded 10% of the initial content of CFC-12. In general, the degradation rate coefficient of CFC-12 was much slower compared to CFC-11, resulting in approximate half-lives of 24 days and 2 days, respectively. Theoretical predictions of the rates for halomethane degradation by reductive dechlorination suggest that the rates for CFC-11 and CFC-12 should be almost equal, but experimental evidence shows that CFC-12 is degraded at rates that are at least 10 times slower than those for CFC-11 (13, 19, 28, 32, 33), a result that was also apparent in this study.
Modeling the Emission and Degradation of Foam Released Halocarbons in Landfills: To evaluate the determined degradation rates in a landfill scenario incorporating all governing processes to the overall fate of the BA, a landfill reactor scenario was set up. It is assumed that the foam waste was disposed of under ideal conditions for microbial degradation of the continuously released BA. Ideal conditions would be to cut the foam panels in moderate sizes, avoid too much compaction, and to mix in anaerobic sludge as inoculum. It was assumed that the foam was cut into pieces(5 cm cubes) and codisposed together with a mixture of organic wastes without any compaction. Compaction was avoided to deteriorate the foam structure as little as possible because any deterioration will enhance the release of BA from the foam. The waste and landfill data were based on typical landfill scenarios as defined in the original MOCLA description. The model is used to evaluate the fate of the degrading BAs (CFC-11, CFC-12, HCFC-22, and HCFC-141b). The HFCs were not shown degradable in the microcosm experiment, and their fates in landfills are, therefore, not modelled.
The degradation rate coefficient, λ, used in MOCLA-FOAM is referring to the water concentration of the compound while the microcosm determined, k1, is referring to the gas concentration. The chosen degradation rate coefficients, k1, in the runs with MOCLA-FOAM were average microcosm rate coefficients ( 0.029 d-1 for CFC-12).
Based on the foam characteristics, the time-dependent release rate was determined using the release model presented in Kjeldsen and Jensen.
Model runs with the defined landfill scenario showed that the leachate and the diffusive loss through the cover soil were negligible fate routes; the release of BA was either released with LFG or being degraded. Since the diffusion coefficient is depending both on the BA and the polymer, different diffusion coefficients were used. A sensitivity analysis on changing diffusion coefficient and degradation rate is performed. The table presents the effect of lower degradation rate coefficient (a factor of 10; 0.1k1) and higher diffusion coefficient (a factor of 10; 10D) to how much BA has been released over a 20-year period, and the percentage of the released amount, which was emitted with gas or degraded. The result shows that CFC-12 and HCFC-22 released from polystyrene had the largest tendency of the four BAs to be emitted with the gas.
The evaluation by using the model MOCLA-FOAM clearly indicates that the emission of BAs disposed of at landfills in foam insulation waste may be attenuated by microbial degradation reactions. However, to which extend the BAs are being attenuated depends especially on how fast a degradation can be obtained under real landfill conditions and also on the release rate of BA from the foam waste. For the current disposal of foam waste in landfills it is difficult to predict the fate of the released BA. In normal landfill operation, waste is compacted to gain landfill volume. How the compaction is affecting initial and long-term release rates of BA is not known. In the future, the use of HFCs as BAs will increase. Since the HFCs seem persistent in landfill environments a larger fraction of the HFC disposed of in foam waste is expected to be emitted to the atmosphere in comparison to the previously used CFCs and HCFCs.

Average First-Order Degradation Rate Coefficients (K₁) and Half-Lives (t_1/2) for Anaerobic Degradation of BAs in Microcosm Experiments Added Different Types of Waste Materials and Inoculated with Digested Sludge from a Biogas Plant Receiving Agricultural Waste^a

Blowing agent		Average degradation rate coefficients					Methane production
		C0^b μg L^-1	K1 d^-1	t½ d	R^2	λ d^-1	T^e d
Waste A: fresh organic waste collected from Danish households – 1. Round
CFC-12	CCl2F2	225	0.029±0.004	23.9	> 0.910	2.90	15
Waste B: Older pre-disposed waste from an American landfill
CFC-12	CCl2F2	-	-	-	-	-	-
Waste C: Waste from a laboratory experimental digester simulating landfill conditions
CFC-12	CCl2F2	-	-	-	-	-	-

^aThe minimum regression coefficient (R²) obtained from fitting the experimental data to a first-order degradation process out of triplicate replicates. The water-based first-order degradation rate coefficients, λ, calculated by eq SI15 and SI16 in the SI are also given.^bAverage initial concentrations; C₀.^eThe average time period; T for production of an amount of methane corresponding to 20% vol in the headspace.

An Analysis Using MOCLA-FOAM for Evaluating Compound Fate Sensitivity to Diffusion Coefficient,D(m²s^-1) and Degradation Rate Coefficient,K₁(d^-1)^a

	Polystyrene foam
	CFC-12		HCFC-22
	0.1K1	K1	0.1K1	K1
Diffusion:D^b
Fraction of initial content release (%)	52	52	100	100
Released amount which has been emitted with gas/degraded (%)	40/60	6/94	57/43	12/88
Diffusion: 10D
Fraction of initial content released (%)	98	98	100	100
Released amount which has been emitted with gas/degraded (%)	40/60	6/94	57/43	12/88

^aThe table also presents the effect of lower degradation rate coefficient (a factor of 10; 0.1K₁) and a higher diffusion coefficient (a factor of 10; 10D).^bDiffusion coefficients used: 5.1 10^-14m²s^-1(CFC-12)(Vo, C. V.; Paquet, A. N. An evaluation of the thermal conductivity of extruded polystyrene foam blown with HFC-134a or HCFC- 142b. J. Cell. Plast. 2004, 40, 205-228); 5.4 10^-12m²s^-1(HCFC-22)(36). The degradation rate coefficients,K₁, in the runs with MOCLA-FOAM are average microcosm rates based on values listed (0.029 and 0.015 for CFC-12 and HCFC-22 respectively).

Conclusions:

Chlorofluorocarbons (CFCs), hydrochlorofluorocarbons (HCFCs), and hydrofluorocarbons (HFCs) have been used as blowing agents (BAs) for foam insulation in home appliances and building materials, which after the end of their useful life are disposed of in landfills. The objective of this project was to evaluate the potential for degradation of BAs in landfills, and to develop a landfill model, which could simulate the fate of BAs in landfills.
Degradation of all studied CFCs and HCFCs was observed regardless the type of waste used. In general, the degradation followed first-order kinetics. The degradation rate coefficient was directly correlated with the number of chlorine atoms attached to the carbon. The highest degradation rate coefficient was obtained for CFC-11, whereas lower rates were seen for HCFC-21 and HCFC-31. Equivalent results were obtained for CFC-12. HCFC-141b was also degraded with rates comparable to HCFC-21 and CFC-12. Anaerobic degradation of the studied HFCs was not observed in any of the experiments within a run time of up to 200 days. T

Executive summary:

Chlorofluorocarbons (CFCs), hydrochlorofluorocarbons (HCFCs), and hydrofluorocarbons (HFCs) have been used as blowing agents (BAs) for foam insulation in home appliances and building materials, which after the end of their useful life are disposed of in landfills. The objective of this project was to evaluate the potential for degradation of BAs in landfills, and to develop a landfill model, which could simulate the fate of BAs in landfills. The investigation was performed by use of anaerobic microcosm studies using different types of organic waste and anaerobic digested sludge as inoculum. The BAs studied were CFC-11, CFC-12, HCFC-141b, HFC-134a, and HFC-245fa. Experiments considering the fate of some of the expected degradations products of CFC-11 and CFC-12 were included like HCFC-21, HCFC-22, HCFC-31, HCFC-32, and HFC-41. Degradation of all studied CFCs and HCFCs was observed regardless the type of waste used. In general, the degradation followed first-order kinetics. The degradation rate coefficient was directly correlated with the number of chlorine atoms attached to the carbon. The highest degradation rate coefficient was obtained for CFC-11, and equivalent results were obtained for CFC-12. Anaerobic degradation of the studied HFCs was not observed in any of the experiments within a run time of up to 200 days. The obtained degradation rate coefficients were used as input for an extended version of an existing landfill fate model incorporating a time dependent BA release from co-disposed foam insulation waste. Predictions with the model indicate that the emission of foam released BAs may be strongly attenuated by microbial degradation reactions. Sensitivity analysis suggests that there is a need for determination of degradation rates under more field realistic scenarios.

Reference 4

Endpoint:

biodegradation in soil

Type of information:

experimental study

Adequacy of study:

supporting study

Study period:

Not specified

Reliability:

2 (reliable with restrictions)

Rationale for reliability incl. deficiencies:

other: Taken from publically available data, and is considered accurate based on the registrants experience of the substance.

Qualifier:

no guideline followed

Principles of method if other than guideline:

Refer to "Detail on Experimental Conditions" listed below.

GLP compliance:

not specified

Test type:

laboratory

Specific details on test material used for the study:

Details on properties of test surrogate or analogue material (migrated information):
Not specified

Radiolabelling:

not specified

Oxygen conditions:

aerobic/anaerobic

Soil classification:

not specified

Soil no.:

#1

Soil type:

loamy sand

% Org. C:

2.1 - 3.2

pH:

> 5.1 - < 6.9

Bulk density (g/cm³):

1.55

Soil no.:

#2

Soil type:

sandy loam

% Org. C:

1.2 - 1.5

pH:

> 5.1 - < 6.9

Soil no.:

#3

Soil type:

other: coarse sand/gravel

% Org. C:

0.3 - 0.6

pH:

> 5.1 - < 6.9

Details on soil characteristics:

Field Site and Soil Sampling: Soil samples were collected at Skellingsted Landfill south of Holbæk, Western Sealand, Denmark. Skellingsted Landfill received a total of approximately 420 000 t of waste between 1971 and 1990. The composition of the waste was approximately 60% municipal solid waste and40%bulky waste, industrial waste, and sewage treatment sludge. The landfill is situated in an abandoned gravel pit located in an area of alluvial sand and gravel sediments. The landfill is uncontrolled with no liners or gas extraction system. The LFG migration has been intensively studied because of a gas explosion accident in 1991. The LFG is mainly migrating horizontally through the sides of the landfill due to the stratified compaction of the waste. The soil was sampled at a test station on the landfill border where the average methane emission was 25 mmol m-2 h-1 (maximal emission was 189mmolm-2 h-1) measured during a 1-yr field campaign. The soil was sampled in 5-cm intervals from the surface to 30 cm depth and in 10-cm intervals from 30 to 90 cm below the surface. Soil samples were collected using a hand auger and stored at 4°C in darkness in closed containers to avoid dehydration. Before storage, the soil was sieved through an8-mmmeshto increase homogeneity. The soil was analyzed for the following parameters: grain size distribution, soil moisture content, organic carbon content, pH, copper content, ammonium, chloride, nitrate, and sulfate. All soil analyses were conducted according to standard methods approved by the Danish EPA.

Soil Characteristics: The soil was characterized and analyzed for different soil parameters as a function of sampling depth prior to column startup. The soil was characterized according to the USDA classification. Three soil layers based on the granulometric composition could be identified: loamy sand (0-35 cm), sandy loam (35-70 cm), and coarse sand/gravel (70-90 cm) (Table 2). The bulk density of the loamy sand layer was 1.55 while the porosity was 0.38. The soil moisture content varied between 9 and 33% w/w with the upper 35 cm being wettest. The soil organic carbon content showed a maximum of 3.2%w/w around 20-cm depth and decreased to 0.3 at 85-cm depth. The total nitrogen content showed a similar pattern with a maximum content of 3.5 g of N (kg of dry soil)-1 at 20-cm depth. The soil pHCaCl2 also showed a maximum (6.9) at 20 cm below the surface and decreased downward to 5.1 at 85 cm. The copper content varied between 2.5 and 8.9 mg of Cu (kg of dry soil)-1 and showed no trend according to depth distribution. The highest chloride concentration (7.5 mg kg-1) was found in the surface soil. Two other minor chloride maxima (3.8 and 3.0 mg (kg of soil)-1) were observed at 20- and 85-cm depth.

Soil No.:

#1

Duration:

3 wk

Soil No.:

#2

Duration:

3 wk

Soil No.:

#3

Duration:

3 wk

Soil No.:

#1

Initial conc.:

25 other: ug L-1

Based on:

test mat.

Soil No.:

#2

Initial conc.:

25 other: ug L-1

Based on:

test mat.

Soil No.:

#3

Initial conc.:

25 other: ug L-1

Based on:

test mat.

Parameter followed for biodegradation estimation:

other: methane and oxygen counter-gradient system.

Details on experimental conditions:

Column Experiments: Column experiments simulating a landfill top cover soil matrix through which gas was transported were carried out. The degradation process was examined in a methane and oxygen counter-gradient system. The columns were packed with landfill cover soil and continuously fed in opposite ends with methane gas (containing trace components) and air. The soil was packed in the columns in the same sequence as sampled in the field. The system consisted of a tube made of rigid PVC, 100 cm long by 8 cmi.d. The PVC tube was closed at both ends with PVC end caps fitted with rubber O-rings to ensure a gastight fit. The PVC cap positioned at the bottom end of the column had one inlet while the PVC cap positioned at the top end of the column had one inlet and one outlet. A perforated plate was located at the bottom of the column so that soil could be packed in the tube.A3-cm layer of sterilized gravel (grain size of 2-3 mm) was placed at the bottom of the column to ensure homogeneous gas distribution. Sampling ports were located along the column length at intervals of 5 cm from the first port, which was positioned 5 cm from the inlet at the bottom. The sampling ports were equipped with Teflon-coated silcone septa, which enabled taking gas samples by a gastight syringe needle. The gas samples (3 mL) were transferred into evacuated glass tubes (Venoject, Terumo Europe n.v., Belgium) and analyzed by gas chromatography. The artificial LFG, which consisted of 50/50% v/v CH4/CO2 was kept in Tedlar bags (SKC Inc., Eighty Four, PA) and fed to the bottom inlet of the column by gastight piston pumps (FMI Lab Pump, model QG, Fluid Metering Inc., Syosset, NY). CFC and HCFC mixtures were added to the CH4/CO2 mixture in the Tedlar bag. The inlet concentrations varied between 15 and 250 μg L-1 depending on the compound, which are within the range of typical LFG concentrations of trace gases. Atmospheric conditions were obtained at the top of the column by passing an air stream through the chamber on top of the soil column (approximately 100 mL min-1). This simulated ambient air over soil cover surface with O2 supply by vertical diffusion into the soil column. Dehydration of the surface soil was avoided by passing the inlet air stream through a washing bottle containing distilled water. The air stream through the chamber on top of the soil column was measured with a soap film flow meter. Gas samples were taken from the Tedlar bags and from the outlet of the columns to control mass balance for the system. Neon was added as conservative tracer during periods of the experimental run. The experiments were carried out at room temperature (22 °C). The average porosity of the packed soil column was 0.52 while the gas filled pore space was 0.22. The porosity of the packed soil column was calculated based on the volume of the column, the soil mass, and the particle density. The experimental setup included three soil columns: two microbial active columns permeated with CFCs and HCFCs, respectively, and a sterilized control column. In the first experimental trial, the degradation was studied under stable conditions with a constant inlet flow. The inlet flow at the bottom of the column was 2.6 mL min-1 corresponding to a gas flux of 0.76 m3 of LFG m-2 d-2 and a methane flux of 250 g m-2 d-2, which is in the mid to high range of reported landfill methane fluxes. Assuming a 20-m-deep layer of waste, this is equivalent to a generation rate of about 13.9 m3 of LFG (m of waste)-3 yr -1, which can be expected within the first 10-15 yr after disposal. The inlet bottom flow was measured by timing the transport of water drops through a defined glass tube inserted between the pump and the column inlet. To obtain steady-state conditions (homogeneous distribution of gas, extinction of sorption capacity of the soil profile) columns were left with gas for 5 d before initial sampling; thereafter, the experiments were run for at least 3 weeks.
To study the degradation processes under more dynamic conditions, column experiments with variable flow were performed. Under natural conditions, the gas flux and the gas gradient system varies, influenced both by changes in the LFG production and by changes in barometric pressures.
The experiment was carried out with CFC-11, CFC-12, HCFC-21, and HCFC-22. Experiments were carried out with eight different inlet flows varying between 0.82 and 14.25 mLmin-1 corresponding to a gas fluxes range of 0.24-4.09m3m-2d-1, which are realistic gas fluxes from the top covers of landfills. Gas fluxes of approximately 0.25 m3m-2d-1 are representative for older landfills or sites with gas collection systems, while new and active landfills with high gas production can have gas fluxes of up to 5 m3m-2d-2. For each flow condition, samples were extracted daily over a period of 5 d. The system was given a period of 2 d between different flow conditions to adjust and reach a steady state. At the end of the experiment, soil samples were taken for analysis of chloride from the column permeated with HCFCs.
The control column was identical to the active columns except that the soil had been sterilized by autoclaving (three times for 1 h each time) and mercury chloride (0.5 g kg-1) had been added thereafter to avoid microbial activity. The control column was run in parallel and operated similar to the active column.
Batch Experiments: To verify observed degradation patterns in the soil columns, simple batch experiments were conducted. Soil was incubated with trace components under both aerobic and anaerobic conditions. A fixed amount of soil (20 g of moist soil) was added to a 117-mL batch container equipped with Mininert (VICI AG, Schenkon, Switzerland) valves made of Teflon. The valves enabled gas to be sampled or injected by a hypodermic needle and a syringe. To simulate landfill conditions, the gas phase in the batch containers was flushed with a 50/50% mixture of methane and carbon dioxide. 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). The soil in the aerobic experiments was sampled at 15-20 cm depth, while the soil used in the anaerobic experiments was sampled at 50-60 cm depth. Gas samples containing the test compound were removed from gaseous stock solutions by a gastight glass syringe and injected into the batch containers. The degradation of the VOCs was studied in single compound tests. The initial concentrations were generally selected in the order that they were in the range of typically trace gas concentrations in LFG (10-250 mg m-3). Gas samples withdrawn from headspace were sampled periodically and analyzed by gas chromatography. The batch experiments were conducted at room temperature (22 °C). All aerobic batch experiments were carried out in series of four, while the anaerobic batch experiments were conducted in duplicate. To check if any disappearance could be due to non microbial processes (abiotic degradation, sorption, and volatilization) control batches with sterilized soil (autoclaving followed by addition of sodium azide (0.2 g kg-1)) were conducted.

Soil No.:

#1

% Degr.:

30

Parameter:

test mat. analysis

Soil No.:

#2

% Degr.:

30

Parameter:

test mat. analysis

Soil No.:

#3

% Degr.:

30

Parameter:

test mat. analysis

Transformation products:

not measured

Details on transformation products:

Not specified

Evaporation of parent compound:

not specified

Volatile metabolites:

not specified

Residues:

not specified

Details on results:

Methane Oxidation and Degradation of Trace Components in a Stable Column System: Methane.
In the control column, the concentration profiles for CH4 and CO2 were almost identical and showed a decrease from 20 cm to the top. The concentration profiles for O2 and N2 show that air was penetrating throughout the whole column. The CH4 concentration profile shows a decrease upward toward the surface, with a maximum decline around 20 cm. Compared to CH4, the upward decrease in the concentration for CO2 is less pronounced, indicating CO2 production. The O2 and N2 concentration profiles show that air is diffusing into the soil matrix from the ambient air. The O2 concentration declines downward with depth and from 35 cm down the column becomes anoxic. The removal of O2 and increasing CO2/CH4 ratio upward in the column indicates methane oxidation. The N2 concentration is much higher in the lower part of the active column as compared to the control column. This is caused by a volume reduction from methane oxidation (3 mol turning into 1 mol) creating an under pressure and thereby enhancing the transport of atmospheric air into soil system. Increasing the supply of O2 into the column will have a positive effect on methane oxidation. The significant effect of the methane oxidation process on the physical gas transport behavior in the column causes the mechanisms controlling the gas flow to be complex, including both advective and diffusive transport, and makes it difficult to compare gas profiles from the active and control column directly. Steady-state gas profiles were obtained within the first 4 d after start-up, indicating that a microbial community of methane oxidizers was already well-established in the soil.
In general, the soil columns showed a high capacity of methane oxidation giving methane oxidation rates between 185 and 210 g m-2 d-1 corresponding to a reduction of 74-81%. The columns showed stable activity, as the removal was fairly constant during the 3-week period that the experiment lasted. These methane oxidation rates are consistent with results reported by Kightley et al., who obtained maximum rates of 166 g m-2d-2 in soil cores of porous coarse sand collected from a landfill site known to emit methane. De Visscher et al also found comparable oxidation capacities of up to 240 g (m of column)-2d-1in columns packed with soil originating from a landfill cover. The lower methane oxidation rates observed in soil columns with HCFCs are probably a result of increased competition between methane and the trace components for the enzyme, due to the higher initial concentration as compared to the soil column with CFCs or due to build up of toxic by products that inhibit the microbial activity. Matheson and co-workers observed irreversible inhibition of methane oxidation by HCFC-21 and HCFC-22. Furthermore, the HCFCs also proved inhibitory to the methanol dehydrogenase (driving the second step in methane oxidation pathway), suggesting that the HCFCs also disrupt other aspects of C1 catabolism in addition to MMO activity. The methane mass balance for the control column was 102% (± 6) recovery indicating no losses and ensuring a tight system. The tracer mass balance showed an average recovery of 101% (± 5) for the control column and an average recovery of 99% (± 8) for the active columns measured over a 10-d period.
CFCs. In the control column, the relative concentration profiles for CFC-11 and CFC-12 are almost identical and show a sharp decrease from 20-cm depth to the top. On the basis of mass balances, no degradation was observed in the control column. CFC-11 and, to a lesser extent, CFC-12 were degraded in the active soil columns. The average degradation capacities for CFC-11 and CFC-12 were 1.0 x 10-2 and 5.8 x 10-3 g m-2 d-1 corresponding to a removal efficiency of 90% and 30%, respectively.
The decline in the CFC-11 concentration profile in the lower part of the column indicates that the removal was due to anaerobic degradation. This was verified by anaerobic batch experiments where CFC-11 was rapidly degraded. Also CFC-12 was degraded in anaerobic batch experiments, however at a lower rate, which is consistent with the column results and with observations of other authors. No degradation of CFC-11 and CFC-12 was observed in aerobic batch experiments despite the fact that rapid methane oxidation occurred with rates up to 112 μg g-1 h-1. Unfortunately, it was not possible to measure degradation products in the soil columns because of much higher detection limits for HCFCs as compared to CFCs.
HCFCs. Average degradation capacities for HCFC-21 and HCFC-22 were 1.1 x 10-1 and 7.6 x 10-2 g m-2d-1, respectively. However, as compared to the CFCs, the degradation of the HCFCs seems to be located in the upper oxic part of the column. For both compounds, the steepest decline in the concentration profile is observed at 20 cm corresponding to the methane oxidation zone with overlapping O2 and CH4 gradients. In aerobic batch experiments incubated with methane, both HCFC-21 and HCFC-22 were rapidly oxidized, and the degradation occurred in parallel with the oxidation of methane. The faster oxidation of HCFC-21 as compared to HCFC-22 is probably a result of the higher stability of HCFC-22 due to the presence of two carbon-fluoride bonds. This is consistent with the higher removal efficiency of HCFC-21 (61%) as compared to HCFC-22 (41%) observed in the soil column. On the contrary, HCFC-21 and HCFC-22 were slowly degraded under anaerobic conditions in anaerobic batches. The degradation of HCFC-22 was very slow, and only 35% degradation was obtained within the duration of the experiment. The degradation of the HCFCs was more than 400 times slower under anaerobic conditions as compared to the oxidative experiments. In a previous study, soil samples from different depths incubated with methane showed that the methane oxidizers were very active in oxidizing HCFCs down to a depth of 50 cm below the surface. Maximum rates were obtained with soil from 15- to 20-cm depth, which is consistent with results obtained in the column experiments. The oxidation rates decreased dramatically at 50-cm depth, indicating that the methane oxidizers were located in the upper part of the soil profile where both methane and oxygen were present.
Methanotrophs expressing the soluble form of MMO (sMMO) were dominating in the upper 20 cm of the soil at Skellingsted Landfill: 15 isolates of type II (where 10 carried the genes for sMMO) were identified and only one type I. This is compatible with the low copper concentration of 4.7 mg g-1 measured in the soil, which corresponds to a copper concentration of 4.7 μg/L-1 in soil water (using a
soil-water distribution coefficient of 1000) as sMMO is only produced in the wild-type organism at very low copper concentrations (<16 μg/L-1). Methanotrophs producing sMMO are particularly valuable in landfill soil covers as the enzyme sMMO has a broad substrate specificity and is known to catalyze faster cometabolic degradation than particulate MMO.
Dehalogenation in the column study was measured by evaluation of the stoichiometric release of chloride by ion chromatographic analysis at the end of the experiment. Approximately 60% of the total HCFC chloride was released and mainly in the upper part of the column. Maximal chloride content (45 mg kg-1) was located at 20-cm depth as compared to the initial concentration of 3.5 mg kg-1. In comparison, the chloride concentration in the lower part of the column (50 cm) was less than 7 mg kg-1.
Methane Oxidation and Degradation of Trace Components in Column Experiments with Variable Inlet Flow: Increasing the inlet flow led to a drop in methane oxidation (i.e., for a flow of 4.1 m3 m-2 d-1 only 24% of the methane was oxidized). Soil gas profiles of methane and oxygen also showed that the oxidation zone was moved upward in the column when the inlet flow was increased. Similarly, the degradation of HCFCs decreased from approximately 75 to 20% when increasing the flow from 0.24 to 4.1 m3 m-2 d-1. In column experiments with CFCs, a drop in degraded mass was also observed when increasing the inlet flow. However, the drop was much less pronounced than for HCFCs, with values falling from 98% to 86% for CFC-11 and from 40% to 10% for CFC-12 when increasing the flow from 0.24 to 4.1 m3 m-2 d-1. Enhanced inlet flow results in diminished retention time and there by a diminished oxidation. However, when increasing the inlet flow, the oxidative zone is moved upward in the column and thereby increases the anaerobic zone. An increased anaerobic zone will enhance the degradation of compounds that are anaerobically degraded and will thereby mitigate the effect of a lower retention time. This again indicates that anaerobic degradation of the CFCs takes place. This experiment clearly demonstrates the complexity of degradation of trace gases in soil covers when both oxidative and reductive processes are in play under variable flow conditions.
Under methanogenic conditions, which exist within the waste, CFC-11 and CFC-12 may undergo reductive dehalogenation leading to accumulation of lesser chlorinated compounds such as HCFC-21, HCFC-31, and HCFC-22, which often are more toxic than the prime compound. However, this study shows that the lower halogenated compounds such as HCFC-21 and HCFC-22 can be rapidly degraded in the upper oxidative zone in landfill soil covers or in the surrounding soil due to their rapid oxidation by the methanotrophic bacteria. The same could be valid for other highly halogenated organic compounds, which are resistant to degradation or only slowly degradable in the aerobic zone, like perchloroethylene, trichloroethanes, etc.

Results with reference substance:

Not specified

Removal Capacities of Methane and Halocarbons Obtained in Soil Column Experiments Permeated with Artificial Landfill Gas^a

Trace gas	Cinlet(μg L^-1)	Methane oxidation		Degradation of halocarbons
		Efficiency (%)	Capacity (g m^-2d^-1)	Efficiency (%)	Capacity (g m^-2d^-1)
CFC-12	25	81±3	210	30±3	5.8 x 10-3

^a50% CH₄and 50% CO₂, v/v. Inlet LFG flow rate = 0.76 m³m^-2d^-1.

 

Maximal Degradation Rates Obtained from Batch Experiments Conducted under Both Anaerobic and Aerobic Conditions^a

Batch experiments		Methane		Halocarbons
Trace gas	Initial gas concn (μg L^-1)	Oxidation rate (μg (g of soil)^-1h^-1)	R^2	Degradation rate (μg (g of soil)^-1 h^-1)	R^2
Aerobic Condition
CFC-12	20	108	>0.983	No degradation
Anaerobic Condition
CFC-12	50			0.0004	>0.937

^aRegression coefficients (R²) obtained from fitting the experimental data to a zero-order oxidation process. The batches held soil water content of 25% w/w and were conducted at room temperature.

Conclusions:

The substance achieved an average removal fo 30% under anaerobic conditions. This indicates that some biodegradation can occur under anaerobic conditions.

Executive summary:

The attenuation of methane and four chlorofluorocarbons was investigated in a dynamic methane and oxygen counter-gradient system simulating a landfill soil cover. Soil was sampled at Skellingsted Landfill, Denmark. The soil columns showed a high capacity of methane oxidation with oxidation rates of 210 g m^-2d^-1corresponding to a removal efficiency of 81%. CFC-11 and to a lesser extent also CFC-12 were degraded in the active soil columns. The average removal efficiency was 90% and 30% for CFC-11 and CFC-12, respectively. Soil gas concentration profiles indicated that the removal was due to anaerobic degradation, which was verified in anaerobic batch experiments where CFC-11 was rapidly degraded. HCFC-21 and HCFC-22 were also degraded in active soil columns (61% and 41%, respectively), but compared to the CFCs, the degradation was located in the upper oxic part of the column with overlapping gradients of methane and oxygen. High oxidation rates of methane and HCFCs were obtained in soil microcosms incubated with methane. When increasing the column inlet flow, the oxidation zone was moved upward in the column, and the removal efficiency of methane and HCFCs decreased. The removal of CFCs was, however, less affected since the anaerobic zone expanded with increasing inlet flow rates. This study demonstrates the complexity of landfill soil cover systems and shows that both anaerobic and aerobic bacteria may play a very important role in reducing the emission of not only methane but also trace components into the atmosphere.

Description of key information

Biodegradation in soil

Key value for chemical safety assessment

Half-life in soil:

36 d

at the temperature of:

20 °C

Additional information

Various assessments of the propensity to biodegrade in general soil and landfil soil was undertaken. Review of the data indicates that CFC-12 will degrade slowly under anaerobic conditions, but not under aerobic conditions. The value proposed above is for degradation under conditions within peat. The overall conclusion is that the substance does not meet the definition of "readily biodegradable" although it will be removed from the soil compartment via anaerobic degradation over time.