Effects of co-disposal of wastes containing organic pollutants with municipal solid waste—a landfill simulation reactor study

Advances in Environmental Research 7 (2003) 949–960

Effects of co-disposal of wastes containing organic pollutants with municipal solid waste—a landfill simulation reactor study ¨ Jorgen Ejlertssona,*, Anna Karlssona, Anders Lagerkvistb, Thomas Hjertbergc, Bo H. Svenssona a

¨ Department of Water and Environmental Studies, Linkoping University, SE-581 83 Linkoping, Sweden Division of Landfill Science and Technology, Lulea˚ University of Technology, SE-971 87 Lulea, ˚ Sweden c ¨ Department of Polymer Technology, Chalmers University of Technology, SE-412 96 Goteborg, Sweden

b

Accepted 17 August 2002

Abstract Different phases of the life cycle of a landfill receiving municipal solid waste (MSW) were monitored in landfill simulation reactors (LSRs) with the aim of investigating the effects of co-disposal of wastes containing organic pollutants (OPs) with MSW. Two LSRs out of four filled with well-characterised MSW received waste materials containing OPs. These included two types of plasticised PVC flooring materials, freon-blown insulation and phosphorus- and nitrogen-based flame-protected materials. Each of the two LSRs was operated under acid fermentative and neutral methanogenic conditions, respectively as were their corresponding controls, i.e. without extra OP. The methanogenic consortia degrading MSW were hampered by the addition of wastes containing OPs, probably due to the presence of Freon R11 and its degradation product, R21. The concentrations of R11 and R21 ranged between 0.1 and 1800 mg my3 depending on the biogas production rate in the OP-amended LSRs. Losses of butylbenzyl- (26%) and bis(2-ethylhexyl)phthalate (15%) from one of two flooring materials was observed, whereas the other remained unaffected. Methanogenic conditions favoured the loss of plasticisers as compared to acidogenic conditions. Total phosphorus was significantly higher in the OP-spiked LSRs, which indicated a transformation of the non-halogenated flame-retardants. 䊚 2002 Elsevier Science Ltd. All rights reserved. Keywords: Acidogenic; Ammonium; Adsorbable organic halogens; Biogas; Landfill; Leachate; Methanogenic; Nitrate; Phosphate; Simulation; Volatile fatty acids; Municipal solid waste

1. Introduction Landfills receive solid waste from municipalities and industries and often sludge from sewage treatment plants. The landfilled waste will undergo transformations through chemical and biological conversion and degradation leading to a development of four typical phases during the ageing of a landfill (Christensen and Kjeldsen, 1989). An initial oxic phase (1) characterised *Corresponding author. Tel.: q46-13-20-83-33; fax: q4613-20-80-32. E-mail address: [email protected] (J. Ejlertsson).

by aerobic degradation, which is followed by an acid fermentation phase (2), where volatile fatty acids (VFAs) and different alcohols may reach molar concentrations. Next is a methane formation phase (3), which is the longest degradation phase of the microbially active landfill. Finally, an oxic phase (4) occurs when the substrates, which may be converted to methane and carbon dioxide (biogas), have been depleted. The length of the different phases will depend on factors such as the size of the landfill, type of refuse landfilled, and climate. In general, the main part of the active life cycle of a landfill takes place when the waste is under anoxic

1093-0191/03/$ - see front matter 䊚 2002 Elsevier Science Ltd. All rights reserved. PII: S 1 0 9 3 - 0 1 9 1 Ž 0 2 . 0 0 0 9 9 - 0

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J. Ejlertsson et al. / Advances in Environmental Research 7 (2003) 949–960

conditions. Therefore, the microbiological degradation of organic materials is due mainly to anaerobic bacteria and their dynamic interaction. The optimal degradation and energy gain for the anaerobic bacteria in a landfill environment, where carbon dioxide is the dominant compound for respiratory activity, take place with methanogens as the most essential group of bacteria. Therefore, anything that hampers the activity of these bacteria will put constraints on the waste mineralisation. The biotic and abiotic transformation processes of different wastes give rise to pools of organic and inorganic compounds in the gaseous and liquid phases. Such compounds may be emitted to the atmosphere or surface and groundwater basins in the drainage area of the landfill. Many of these compounds are organic pollutants (OPs) including aromatic compounds such as ¨ phenols, phthalates, creosols (Oman and Hynning, 1993), and halogenated hydrocarbons like freons, diand trichloroethene, and vinylchloride (VC) (Willumsen et al., 1988; Laugwitz, 1990). Most of these compounds are of anthropogenic origin and are brought to the landfill as components of building refuse, plastics, paints, materials treated with flame-retardants, cryogenic media, isolation materials, pesticides, and solvents. The formation, appearance, and disappearance of a specific pollutant in a landfill depends on several factors, including the chemical and physical properties of the compound itself, e.g. its chemical reactivity, water solubility, sorption capacity, volatility, and biodegradability. Digestion experiments with municipal waste in pilot-scale reactors, supplemented with different OPs (Pohland, 1991), indicate that these compound-specific properties interact with the physical and chemical conditions prevailing in a landfill, which include temperature, moisture, pH, types of waste, and the degradation capability of the microflora. Hence, it is likely that the landfill development phases have different potentials for the transformation of OPs. In a society based on ‘recycling and reuse’, landfills will still play an important role in the process of waste handling. Thus, the managing techniques to minimise the effects of landfills on the environment will be important. It is also crucial to know the behaviour of different waste materials containing OPs in landfills as well as their possible effects on the overall function of the landfill, i.e., the activity of its bacterial population. The overall aim of the present investigation has been to study the effects of co-disposal of wastes containing OPs with municipal solid waste (MSW). The addition of such wastes will give rise to a number of chemicals in the aqueous and gaseous phases due to leaching and biological transformations. Thus, this approach differs from most other studies where investigated chemicals were spiked as such in the waste. The following questions were formulated: (1) Will the development phases of landfilled waste impact the behaviour of the OP-

containing waste materials due to their inherent difference in leaching characteristics and microbial activity? (2) Will the OP-containing waste materials give rise to compounds that can be traced in the leachate and gaseous phase? (3) Will microbial populations able to attenuate different OPs develop during the ageing process of landfilled waste? (4) Will the addition of wastes containing OPs hamper the general degradation processes of landfilled MSW to biogas? 2. Experimental section To answer the questions listed above, four 100-l landfill simulation reactors (LSRs) (LiU1–4) were started in February 1999. Two were run under methanogenic and two under acidogenic conditions. All reactors were filled with 35–40 kg of characterised MSW (see below); water was added to give a moisture content of 65% (wet weight). The design of the LSRs build on the construction by Stegmann (1981) as modified by Lagerkvist and Chen (1993). One reactor of each to become methanogenic and acidogenic, respectively, was supplied with extra waste materials: two PVC-flooring materials (FLO2 and FLO3), freon-blown insulation, a seat cushion, a radio, and flame-protected clothing materials (Table 1). The rationale behind this was to add materials that could release OPs like freons, different additives in plastics and flame-retardants into the waste mixture and leachate. The amount of R11 added to LiU2 and LiU4 with the insulation material was estimated to be approximately 16 g. The estimation was based on the assumption that a typical insulation material used before 1989 contained 10–12% by weight R11 (Kuhn and Schindler, 1987). The diffusion of R11 from the insulation during use was assumed to be negligible, based on the findings made by Gaarenstroom et al. (1989). Furthermore, the losses of R11 during homogenisation did not likely exceed 10%, since at most 10% of the insulation pores may have been disrupted during the crushing and a major portion of R11 in the pores is dissolved in the polyurethane polymer (Gaarenstroom et al., 1989). A conservative estimate of the amount of R11 added to LiU2 and 4 with the insulation material would then be 8–9% by weight, which in total amounts to 16–18 g of R11. 2.1. Start-up of the LSRs The waste materials originated from a housing area ¨ (Lulea, in Porson ˚ Sweden). Approximately 55 waste sacks (f500 kg wet weight) were collected during 2 days in late January 1994. The waste was sorted by hand into 9 fractions (Table 2). The different fractions (except for half of the metal and the entire noncombustible fraction) were then milled separately in a hammer mill (1–2 cm particle size). After blending the

J. Ejlertsson et al. / Advances in Environmental Research 7 (2003) 949–960

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Table 1 Total amounts of extra waste material added to the LSRs LiU2 and LiU4, respectively Material

Weight added (g)

Target compounds

Amount added (g)

FLO2a

200

BBP DEHP PVC polymer

14 38 114

FLO3a

800

DEHP: PVC polymer

136 384

350

Tetrabromobisphenol A

n.d.

200

Melamine N:

29

Pyrovatex P: N:

36 10

Proban P: N:

38 17

Freon R11

16

Radioa Seat cushion Curtain

b

b

Overallb

Insulationc Sum

750

750

200 3250

a

Hammer milled. Both FLO2 and FLO3 contained fillers (mainly CaCO3 ), pigments, and stabilisers (CayZn soap and epoxidised soybean oil). BBP, butylbenzyl phthalate, DEHP, bis(2-ethylhexyl)phthalate. b Knife milled. The flame-retardants used in the cloth contained Proban and Pyrovatex, which are phosphorus- and nitrogen-based (Sakai, 1987), and the seat cushion contained melamine (1,3,5-triazine-2,4,6-triamine). c Crushed by hand.

milled fractions, the waste was transferred into 60 polyethylene plastic bags in 5 kg ("0.1 kg) portions. The total solid content of the waste was determined to be 64%. LiU1 and LiU3 were each filled with MSW from 8 randomly picked waste portions (in total 40 kg).

For each of LiU2 and LiU4, 7 bags were selected (in total 35 kg). The additional OP materials (Table 1) were divided into smaller portions and separately mixed into every 5-kg waste portion before transfer to the LSRs during the filling of LiU2 and 4. All LSRs were

Table 2 Composition of the MSW used in the study Waste fraction

Wet weight (kg)

Compost fraction Recyclable paper

252.0 65.0

52.9 13.6

Plastic fraction Polyethyleneypropylene Polystyrene PVC Other plastics

31.2 25.8 3.1 0.9 1.4

6.5 5.4 0.6 0.2 0.3

Rubberyleather Combustible Textiles Non-combustible Glass Metals

0.1 66.2 4.5 11.7 35.2 10.8

0.03 13.9 0.9 2.5 7.4 2.3

Sum

476.7

Share of wet weight (%)

100.0

¨ during late January 1994. It was collected from a housing area in Porson

Waste fraction content Diapers, food, moist papers, shells, flower soil, etc. Newspaper, envelopes, books, magazines, cardboard. Plastic bags, toys, plastic foil, containers, etc. Containers, toys, etc. Cable, containers, maps, gloves, vinyl disks, etc. Bags, sponges Shoes Milk containers, mixed plasticymetal materials, etc. Nylon and cotton socks, T-shirts, wool gloves, etc. Batteries, stones, spray cans, medicine containers Cans, jars, etc. Beer cans, tin jars, screw caps, etc.

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closed with a steel lid and tightened with screws. Water was added via the recirculation system to give a moisture content of 65% (wet weight: 33 l to LiU1 and LiU3 and 35 l to LiU2 and LiU4) in the LSRs. 2.2. LSR operation After packing and closure, the two LSRs to be run methanogenically (LiU1 and LiU2) were aerated (1 l miny1 of moisturised air) periodically until day 117 in order to pre-compost the waste. By pre-composting, the acidogenic stage was avoided, allowing methanogenesis to develop as soon as the air was switched off. Thereafter, LiU1 and LiU2 were running methanogenically throughout the experimental period (February 1994 to July 1999). In order to mimic the dilution of leachate that occurs in a landfill due to rainfall and infiltration, an exchange of leachates with tap water was applied. Soon after the onset of stable methanogenic conditions in LiU1 and LiU2 (day 129), 3 l of leachate was withdrawn weekly from all four LSRs and replaced with tap water. This was done for 33 weeks (until day 362). In total 100 l of leachate was taken out and replaced. The replacement of leachate resulted in a tenfold dilution of the leachate, which also was the dilution factor calculated from the decrease in total halides during the same time period (Figs. 1 and 2). After more than 2.5 years of acidic operation (day 1007), LiU3 and LiU4 were slowly turned into methanogenic conditions. On a weekly basis, one litre of leachate was pumped out from the four LSRs into 1-l measuring cylinders. The acidic leachate from LiU3 was then pumped into LiU1, and the methanogenic leachate from LiU1 was pumped into LiU3. The same procedure was undertaken with the spiked LSRs, LiU2 and LiU4. The leachate exchange was done until stable biogas production was established in LiU3 and 4. The recirculation of leachates would to some extent mimic conditions likely to occur in landfill, i.e., acid leachates may reach methanogenic areas and vice versa. During the LSR performance, the gas production was measured continuously. Liquid samples for VFAs, NHq 4 , total phosphorus (Tot-P), and AOX were taken regularly. 2.3. Analytical procedures Gas production was measured by weekly readings of gas meters, which counted every 60 ml of gas produced. Methane was analysed by gas chromatography (GC) and quantified with flame ionisation detection (FID) as ¨ described in Orlygsson et al. (1993). Freon R11 and VC in the biogas produced from the LSRs was analysed by GC and quantified simultaneously with FID and electron capture detection (Ejlertsson et al., 1996a). GC–FID was also used for the quantification of ethanol and VFAs from acetate to caproate. Liquid

Fig. 1. Biogas formation and changes in leachate pH, TOC and total halides (Fy, Cly, Bry) during operation of the methanogenic LSRs: LiU1 (h) and LiU2 (j). Note that the first 200 days of LSR operation is shown for pH.

samples (1 ml) were centrifuged for 20 min at 6000=g. The supernatants were then acidified with a formic acid solution (final concentration of 2 M) containing cyclohexanone as the internal standard. The samples were injected (1 ml) by an auto sampler in an HP 5880 with a Chrompack WCOT fused-silica column (25=0.32 mm2) coated with CP-Sil5CB. Helium was used as a carrier gas at a flow rate of 1.5 ml miny1 and nitrogen was used as a make-up gas at 30 ml miny1. Detector gases were hydrogen (33 ml miny1 ) and air (400 ml miny1). The injector and detector temperatures used

J. Ejlertsson et al. / Advances in Environmental Research 7 (2003) 949–960

Fig. 2. Biogas formation and changes in leachate pH, TOC and total halides (Fy, Cly, Bry) during operation of the initially acidic LSRs: LiU3 (s) and LiU4 (d).

were both at 250 8C. A temperature gradient was used to obtain separation. The initial oven temperature was 60 8C (4 min); the temperature was then raised in steps of 3 8C miny1 to a final temperature of 180 8C. The TOC analysis of the leachates was performed on a Shimadzu TOC-5000 analyser with calibration curves ranging from 0 to 100 mgC ly1. Samples containing more TOC were diluted. The leachate water was not filtrated prior the TOC analysis. Liquid samples for AOX were analysed according to the German standard DIN 38409 on a Euroglas microcoulometer. Inorganic halides (Cly, Bry, and Iy) were

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analysed by microcoulometric titration according to the German standard DIN 38409. Ammonia and orto-phosphate were separately analysed on unthawed and untreated leachate samples using a FIAstar 5020 analyser (Tecator) equipped with an FIAstar 5102–001 injector and a 5022 spectrophotometer for detection. Calibration curves for ammonia were made between 1 and 10 mgN ly1 (100 ml injected) and 0.25 and 5 mgP ly1 for orto-phosphate (30 ml injected). Samples with higher ammonia or orto-phosphate concentrations were diluted in Millipore water. Tot-P was treated with peroxodisulphate according to SS 028127 and determined using the same methodology as for orto-phosphate. Dissolved phosphorous was determined in filtrated (pore size 0.45 mm; Millipore) and peroxodisulphate-treated leachate using a FIAstar (see above). Particulate-bound phosphorous was determined as the difference between Tot-P and dissolved phosphorus. Nitrogen quantification of the N-containing materials was done at the Department of Soil Science, SLU (Sweden). Phosphorus content in the P-containing materials was determined using ‘the oxygen flask method’ (MacDonald, 1961). Determination of molecular weight distributions (MWD) and the content of plasticisers in the two PVC floorings investigated was done using size exclusion ¨ chromatography as described by Abbas ˚ and Sorvik (1975). A Waters 150 CV equipped with two mStyragel columns was operated at 25 8C with tetrahydrofurane (THF) as the solvent. The concentration of the PVC sample solution was 3 g ly1 and the solutions were heat-treated at 120 8C for 3 h in order to dissolve molecular aggregates in the PVC flooring material, after which 300 ml were injected into the columns. The MWD and molecular weight averages were obtained via universal calibration from narrow polystyrene standards using software supplied by Waters. The amount of plasticisers remaining in both flooring materials was determined by comparing the peak heights of the plasticisers in the LSR samples with an untreated reference sample using the PVC polymer as the internal standard. Differences in the amount of plasticiser between sample and reference that are greater than 2% are statistically significant with the methodology undertaken. 3. Results and discussion 3.1. Degradation of MSW in the LSRs LiU1 and LiU2 were aerated periodically until day 117, when the latter showed a continuous biogas formation and the pH had stabilised at about neutral (Fig. 1). After this, the gas production as well as the methane content increased in both LiU1 and 2. The methane concentration in the biogas produced was constantly

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50–55% throughout the experimental period (data not shown). Unfortunately, a small leak in the gas collection system for LiU1 made it impossible to follow the gas production in this lysimeter until the leak was found and sealed. LiU3 and LiU4 produced large amounts of VFA and became acidogenic after a few days of LSR performance (Figs. 2 and 3). During the first 30–40 days, approximately 500 l of gas was produced and in the same period, pH stabilised approximately 5.5. From day 200 until day 1000, an increase in pH to 5.8 was observed for LiU4, whereas LiU3 remained at pH 5.5 (Fig. 2). At day 1007, the leachate exchange was started between the acidic and methanogenic LSRs and after a few weeks LiU3 was producing biogas at a high rate (Fig. 2). After 4 months of leachate exchange, the pH in LiU3 was neutral. At this time, the leachate from LiU4 had a pH of approximately 6. The leachate exchange between LiU1 and LiU3 was stopped after approximately 6 months (day 1186). After another year (fday 1500), LiU4 also had reached a pH of 7 and was producing biogas at high rates. The onset of biogas production observed for the acidic LSRs was also reflected in the pH and TOC content (Fig. 2). The TOC dropped to approximately 500–300 mgC ly1 approximately day 1100 for LiU3 and between day 1400 and 1500 for LiU4, which coincided with the increased gas production. For LiU1 and 2, acetate was the dominating VFA, peaking at a concentration of almost 200 mM after 50– 100 days of LSR operation (Fig. 3). The highest levels of fermentation products were observed in LiU2, resulting in a longer initial period of low pH than was the case for LiU1. Approximately 200 days after start-up, all VFA were diluted andyor transformed to levels below the detection limit (0.05 mM). Within 2 weeks of LSR performance, fermentation product concentrations as high as almost 0.5 M were observed in LiU3 and LiU4. The acetate concentration was lower in LiU4 (75–100 mM) than in LiU3 (f150 mM). The levels of VFA remained at these high concentrations in both acidic LSRs until day 129 when leachate dilution with water started. The dilution procedure caused the total concentration of VFA to drop from approximately 500 to 70 mM (day 362), with acetate, butyrate, and caproate as the dominating compounds (25, 25, and 10 mM, respectively). The washout of VFA occurred in a similar way to the dilution of inorganic halides (Figs. 2 and 3). Ethanol was, however, decreasing faster than the dilution rate. Thus, ethanol was further metabolised during this initial period, while the VFA pool was only diluted. Once stopped, the degradation of organic matter did not resume until the conditions became neutral due to the onset of methanogenesis, even though considerable amounts of fermentative substrates were available and the VFA concentration was decreased by one order of magnitude. A similar observation was made for a biogas

Fig. 3. Development of VFA in LiU1 and 2 during the first 200 days of LSR performance and occurrence of VFA in LiU3 and 4 during the LSRs performance. VFA was not detected in LiU1-2 after 200 days of LSR performance and was not observed in LSR 3–4 after 1600 days of performance: Acetate (s), propionate (⽧), butyrate (m), ethanol (j) and caproate («).

reactor in which silage was used as substrate in a twostep process (Nordberg, 1996). These authors observed that only a small fraction of the silage was further fermented in the acidic reactor and that it was rather the VFAs already present in the silage that gave rise to methane in the second reactor.

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ammonium concentration in LiU3 resembles a washout curve, but since the leachate was exchanged with LiU1, which held a concentration of approximately 400 mg ly1, a consumption of ammonium must have occurred in LiU3 and later on also in LiU4. The disappearance of ammonia coincided with the onset of high gas production in both LiU3 and LiU4 and is thus probably due to the fixation of nitrogen into biomass. The decrease in the biogas production rate in LiU3 occurred when ammonia was depleted, indicating a nitrogen limitation in the system.

Fig. 4. Changes in leachate AOX, ammonia (NHq 4 ), total phosphor (Total P) and orto-phosphate during operation of the methanogenic LSRs: LiU1 (h) and LiU2 (j).

During the LSR performance, the ammonia concentration decreased under the dilution period to approximately 100 mg ly1 in LiU1 and to approximately 200 mg ly1 in LiU2. The ammonia concentration then increased to between 400 and 450 mg ly1 in both LSRs (Fig. 4). In LiU3, it levelled off at approximately 500 mg ly1, and then decreased to zero between days 1150 and 1350 (Fig. 5). The ammonium concentration in LiU4 fluctuated but decreased until day 600 when it was approximately 150 mg ly1. The concentration then increased to the former stable level of LiU3, but decreased again approximately day 1450 until it reached approximately 150 mg ly1 at day 1875. The decrease in

Fig. 5. Changes in leachate AOX, ammonia (NHq 4 ), total phosphor (Total P) and orto-phosphate (Orto P) during operation of the initially acidic LSRs: LiU3 (s) and LiU4 (d).

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Fig. 6. Normalised MWD of FLO2 and FLO3 taken after 1640 days of LSR operation: Reference sample (—), LiU2 («) and LiU4 (----).

3.2. The fate of specific OP-containing wastes during LSR performance LiU2 and LiU4 were opened and sampled for PVC plastics on days 1122 and 1640. The PVC samples were examined for changes in the PVC polymers, i.e., MWD, and remaining amounts of plasticisers. For FLO3, no difference in the MWD could be observed between aged samples and the untreated reference (Fig. 6). In the case of FLO2, the MWD aged in LiU4 seems to differ somewhat from that of the reference and LiU2 (Fig. 6). However, the averages of the molecular weights are all within the experimental errors, and it can be concluded that polymer chains have not been affected during the lysimeter operation. In the debate regarding PVC plastics, the public has been concerned about transformation of the PVC polymer into VC in landfills, since the landfill gas is known to contain VC (Willumsen et al., 1988; Laugwitz, 1990). However, the analysis of the

MWD does not show any indications of a depolymerisation process of PVC, i.e., formation of VC. Neither have VCs been detected in the biogas from the LSRs (detection limit 0.1 mgVC my3 ). From the MWD in the LSR-treated samples, no increase in the high molecular weight tail was observed, which indicates the absence of any cross-linking reactions between PVC polymers. The peak of the distribution has not been shifted towards lower molecular weights, which shows that no random chain scission reaction has been active. Considering the temperature conditions prevailing in the LSRs and landfills, this is what could be expected. HCl is the dominating compound of low molecular weight formed during degradation of PVC polymers at a temperature relevant for processing, i.e., 150–220 8C. Benzene has been observed at these temperatures, but the amount is in the order of 2% of the amount of HCl formed (Kelen, 1978). Somewhat larger amounts of different aromatics can be found during degradation of

J. Ejlertsson et al. / Advances in Environmental Research 7 (2003) 949–960 Table 3 Plasticiser content in the FLO2 reference and after 1122 and 1640 days of incubation in LiU2 and 4 LiU2

LiU4

BBP (g)

DEHP (g)

BBP (g)

DEHP (g)

Reference Cross-section

14

38

14

38

Day 1122 Cross-section Total loss

13.1 0.9

37.1 0.9

14 0

37.9 0.1

Day 1640 Cross-section Total loss

10.4 3.6

36.5 1.5

12.7 1.3

32.3 5.7

The estimated total loss from FLO2 is based on the assumption that all flooring material added to the LSRs was affected in the same way as the analysed samples. No loss of DEHP was observed in FLO3.

PVC polymers at temperatures relevant for pyrolysis (350–900 8C) (Montaudo and Puglisi, 1991). Almost all of these compounds detected at high temperatures are non-chlorinated and the formation of low-molecularweight organic compounds during degradation of PVC takes place only after the release of HCl. One study (Christmann et al., 1989) reports on trace amounts of VC found during pyrolysis experiments. The sequential degradation of landfilled chlorinated ethylenes formerly used in solvents and cleaning agents may be one likely explanation for the VC found in landfill gas. Reductive transformation of perchloroethylene with formation of VC as intermediate or final products have recently been shown to be performed by organisms in landfill leachate (Kromann et al., 1998) as well as in landfilled waste (Svensson et al., in preparation). A loss of plasticisers from the FLO2 occurred in samples from both LiU2 and LiU4 (Table 3). At sampling day 1122, losses corresponding to 6% BBP and 2% DEHP were observed in samples taken from LiU2. Samples from the (at the time) acidic LiU4 showed no significant loss of plasticisers. The samples taken at day 1640 showed in total a loss of BBP (26%) and DEHP (4%) in LiU2. In total, 3.6 g of BBP and 1.5 g of DEHP were lost from the FLO2 during 1640 days of LiU2 performance. A significant loss of both BBP and DEHP also occurred in the FLO2 samples taken from LiU4 at the later sampling occasion when the LSR produced biogas. The total loss amounted to 9% BBP (1.3 g) and 15% DEHP (5.7 g). The findings show that the major loss of BBP and DEHP from FLO2 occurred in an active methanogenic environment, in contrast to the extractive characteristics of an acidic environment with high TOC values and VFA concentrations as high as 0.5 M. That transformation of different phthalic acid esters (PAEs) by landfill microorganisms

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has been shown previously (Painter and Jones, 1990; Reinhart and Pohland, 1991; Ejlertsson et al., 1996a; Ejlertsson, 1997). One study (Ejlertsson et al., 1996b) showed that the transformation of PAEs might give rise to phthalic monoesters (PMEs), which may accumulate and possibly persist biodegradation. However, (Jonsson et al., in press) found that PMEs occurred in the leachates of all the LSRs. This clearly shows that PAEs have been extracted and transformed in LiU1–4. No loss of DEHP was observed from FLO3 in any of the samples taken from LiU2 and 4. It should be noted that the loss of plasticiser will change the PVC floorings from flexible to rigid (glass transition) which is associated with drastic changes in the diffusion rate of the remaining plasticiser(s). Depending on the size of the plasticiser in question, the permeability below the glass transition is approximately 7–8 orders of magnitude lower than above the glass transition (Chern et al., 1989). For a given PVC formulation, the temperature for glass transition increases when the concentration of plasticisers decreases (Sears and Touchette, 1989) and eventually the glass transition will occur at the temperature of the surrounding media. Thus, the loss of BBP and DEHP from FLO2 during the LSR operation, which all are obtained from conditions above the glass transition, can not be used to estimate the time for the release of all plasticiser(s) present in the plasticised PVC flooring. The elevated amounts of Tot-P in both LiU2 and LiU4 indicate that the P-containing flame-retardants were subjected to transformation (Figs. 4 and 5). Approximately 70% of the phosphorus detected were water soluble, whereas the main part of the rest was particle bound in these two LSRs (Figs. 4 and 5). In LiU1 and 3, the concentration of soluble Tot-P declined to below the detection limit approximately day 1000. The particle-bound phosphorus ranged from below the detection limit to 10 mg ly1 (data not shown). The roughly equal levels of orto-phosphate in the LiU1 and 3 compared to the spiked LSRs indicate that complete mineralisation of the dissolved phosphorous compounds did not occur. Proban forms a large polymer on and inside the cotton fibres (Sakai, 1987), while the Pyrovatex reacts with hydroxyl groups present on the cotton cellulose. Thus, Pyrovatex may be subjected to hydrolysis and set free during cellulose degradation. It may undergo further mineralisation, with a corresponding increase in Tot-P. The excess phosphorus from LiU2 when compared with LiU1 is approximately 50 mg ly1, which equals 1.6 mM of phosphorus. Every carbon connected to such a concentration of phosphorus would give rise to an increase in the TOC levels of approximately 20 mg ly1. If we assume that the dissolved phosphorous is contained within organic compounds originating from the Pyrovatex, then they are most likely small molecules (1–2 carbons), since the TOC

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values in LiU1 and 2 were similar at 200–300 mg ly1. Nitrogen that would be released during degradation of mainly Pyrovatex would not be possible to distinguish from the high background of ammonia, which made up more than 90% of total nitrogen. On day 2076, gas samples from all four LSRs were investigated for the presence of R11 and R21 (dichlorofluoromethane), which is an intermediate that may occur during degradation of R11 by landfill methanogenic consortia (Svensson et al., in preparation). Even after this long period of LSR operation, both R11 and R21 were observed in both LiU2 and 4. As expected, none of these compounds were traced to LiU1 and 3. The concentrations of R11 and R21 were highest in LiU2 and amounted to 10 and 1800 mg my3, respectively. The concentrations of both R11 and R21 were lower in LiU4 (0.1 and 40 mg my3, respectively), which most likely was due to the high gas-production rate in LiU4 as compared to LiU2. LiU4 produced 10–15 l of biogas per week at the gas-sampling occasion, whereas the gas production in LiU2 was approximately 100 ml per week. One study (Ejlertsson et al., 1996a) reported a complete transformation of R11 to R31 via R21 by an LSR inocula. Others (Deipser and Stegmann, 1994) observed that R11, added directly to an LSR, was transformed to R21 during co-disposal with mixed household waste. According to these studies, R11 was transformed under both acidic and methanogenic conditions, with R21 as the sole end product formed. The concentration of R21 observed in LiU2 (1800 mg my3) roughly corresponds to 1% of added R11 and highlights the slow diffusion of R11 from the insulation material (cf. Gaarenstroom et al., 1989). The results also imply that gas from old and mature landfills with low gas production may contain significant amounts of both R21 and R11. Furthermore, in landfills with high gas production, R11 and its degradation product R21 will be diluted with formed biogas, since the diffusion of R11 from the insulation is low. One study (Laugwitz, 1990) investigated 20 German landfills for the occurrence of chlorinated aliphatics and found that both R11 and R21 was observed in more than 80% of the landfills at 1– 10 mg my3 for both gases. In contrast to the results presented here, the study (Laugwitz, 1990) observed decreasing concentrations for both R11 and R21 with landfill age, although gas production rates were not determined and landfills older than 35 years were not included in the study. It is therefore plausible that the biogas in old and mature landfills may contain elevated concentrations of different halogenated compounds that are limited in their diffusion from polymeric material such as R11. The waste materials containing OP gave rise to a small but significantly higher content of halogenated compounds in the leachate, which were measured as AOX in LiU2, compared to LiU1 (Fig. 4). After 1000

days of methanogenic performance, the AOX concentration levelled at 0.40 mg ly1 in LiU1 and 0.55 mg ly1 in LiU2. The amount of ‘normal’ waste was 35 kg in LiU2 compared to 40 kg in LiU1. Thus, it seems that the extra waste materials added to LiU2 gave rise to a higher level of AOX in the leachate. However, the difference could also be because of a lower degree of AOX transformation in LiU2. The AOX content in the leachate increased in the former acidic LSRs when the conditions changed to methanogenic, although a difference in the AOX content in LiU4 compared to LiU3 could not be distinguished (Fig. 5). The higher amounts of AOX detected in LiU3 and 4 could have at least two explanations: 1. Since higher amounts of AOX were initially detected in both LiU1 and 2 (f4 mg ly1) than in LiU3 and LiU4 (f2.5 mg ly1), it is plausible that a larger amount of the AOX was either diluted or degraded during the dilution period (days 117–362). 2. The AOX may have been trapped in different waste materials and was released only after degradation of the waste, i.e., when the conditions became methanogenic and most of the waste degradation occurred. 3.3. General effects of the OP-containing waste on the LSR performance The addition of extra waste material that contained OPs affected the biological degradation of the ‘normal’ MSW. The effects were first observed during the initial aeration period of LiU1 and LiU2, where LiU2 needed a longer aeration period than LiU1 to overcome the acidic phase (Fig. 1). A hampering effect of the OPcontaining waste materials was also shown during the transition of LiU3 and LiU4 into methanogenic conditions. Not only did LiU4 start biogas production over one year later than LiU3, but the initial gas accumulation rate was slower and levelled off at a lower level of mineralisation (Fig. 2). Likely, methanogenic archae in LiU4 were hampered, which is supported by the retarded degradation of TOC, VFA, and ethanol (cf. Figs. 2 and 3). It is also possible that fermenting bacteria were affected by the added OP-containing materials, since the amount of acetate formed was significantly lower in LiU4 than in LiU3 after approximately 30–40 days of LSR performance. A plausible explanation of the negative effects observed is given by the degradation experiments performed on waste samples from the LSRs. These showed that Freon R11 and the degradation product R21 gave rise to a strong inhibition of the methanogenic populations (Ejlertsson et al., 1996a; Svensson et al., in preparation). The other OPs from the added waste materials, which included BBP, DEHP, melamine, Pyrovatex, and PVC-polymers, did not give rise to any observable inhibitory effects on the microbial

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populations for any of the LSRs. Tetrabromobisphenol A and Proban were not included in those studies. However, synergistic effects of these compounds were not studied and may also serve as a possible explanation to the observed inhibitory effects on the methanogenic populations. 4. Conclusions The overall aim of the present investigation has been to study the effects of co-disposal of wastes containing OPs with MSW. This approach differs from most other studies where investigated chemicals were spiked as such in the waste. The results obtained in the study showed that OPs released from the waste had effects on the microbial processes and that the methanogenic consortia degrading MSW were hampered, probably due to the presence of Freons. Also, losses of butylbenzyl(26%) and bis(2-ethylhexyl) phthalate (15%) from flooring materials was observed. The conditions of the methanogenic phase favoured the loss of plasticisers as compared to acidogenic phase. Thus, the landfill development phases certainly affected the behaviour of the OPs. Acknowledgments Ulla-Britt Uvemo (Lulea˚ University) and Dzeneta ¨ Nezirevic (Linkoping University) assisted with the LSR performance and the analysis of the huge amount of samples generated within the project. Lena Lundman ¨ and Maritha Horsing analysed for phosphorus and Morgan Zaar (SLU) for nitrogen in the clothing material. The following persons and companies contributed to the study of the OP-containing waste materials and information about the products: Roland Karlsson (Tarkett ¨ Ronneby, Sweden), Maria Stavasen (Boras ˚ ˚ Vafverier, Boras, ˚ Sweden), Anders Selin (Electrolux, Stockholm, Sweden), Philips (Stockholm, Sweden), and IKEA ¨ (Almhult, Sweden). This project was funded with grants from the Swedish Environmental Protection Board, contract no. 802–335–903-Fr, and the organisations supporting the project on ‘Long term behaviour of PVC-products under soil-buried and landfill conditions’: ECPI, ECVM, ELSA, ORTEP and Norsk Hydro. References ¨ Abbas, J., 1975. On the thermal degradation of ˚ K.B., Sorvik, poly(vinyl chloride). III. Structural changes during degradation in nitrogen. J. Appl. Polym. Sci. 19, 2991–3006. Chern, R.T., Koros, H.B., Hopfenberg, H.B., Stannet, V.T., 1989. Materials selection for gas separation membranes. ACS Symp. Ser. 269, 25–46. Christensen, T.H., Kjeldsen, P., 1989. Basic biochemical processes in landfills. In: Christensen, T.H., Cossu, R., Steg-

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