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Waste Management 28 (2008) 2344–2354 www.elsevier.com/locate/wasman
The decomposition of wood products in landfills in Sydney, Australia F.A. Ximenes a,b,*, W.D. Gardner c, A.L. Cowie a,b a
Forest Resources Research, NSW Department of Primary Industries, P.O. BOX 100, Beecroft, NSW 2119, Australia b Cooperative Research Centre for Greenhouse Accounting, Australia c Forest Resources Research, NSW Department of Primary Industries, 33 Galahad Crescent, Castle Hill 2154, Australia Accepted 9 November 2007 Available online 4 January 2008
Abstract Three landfill sites that had been closed for 19, 29 and 46 years and had been operated under different management systems were excavated in Sydney. The mean moisture content of the wood samples ranged from 41.6% to 66.8%. The wood products recovered were identified to species, and their carbon, cellulose, hemicellulose and lignin concentration were determined and compared to those of matched samples of the same species. No significant loss of dry mass was measured in wood products buried for 19 and 29 years, but where refuse had been buried for 46 years, the measured loss of carbon (as a percentage of dry biomass) was 8.7% for hardwoods and 9.1% for softwoods, equating to 18% and 17% of their original carbon content, respectively. The results indicate that published decomposition factors based on laboratory research significantly overestimate the decomposition of wood products in landfill. Ó 2007 Elsevier Ltd. All rights reserved.
1. Introduction Approximately 8.9 million tonnes of putrescible waste were disposed of in landfills in Australia in 2002/2003 (Productivity Commission, 2006). According to the Australian Greenhouse Office (AGO, 2006), the main components of waste disposed of in municipal solid waste (MSW) landfills in Australia in 2005 were paper and textiles (30%), food (15%), garden and green (15%), wood (3%) and others (36%). Landfills represent a major disposal option for wood and paper products in many countries. It is estimated that approximately 2.3 million tonnes of solid wood products and 2 million tonnes of paper products are placed in all Australian landfills each year (Ecorecycle Victoria, 2004; AGO, 2004). The decomposition of wood products in landfills may be a significant source of greenhouse gas emissions; the anaerobic decomposition of organic components in landfills results in the generation of methane and carbon dioxide in approximately equal proportions (IPCC, * Corresponding author. Address: Forest Resources Research, NSW Department of Primary Industries, P.O. BOX 100, Beecroft, NSW 2119, Australia. Tel.: +61 2 9872 0143; fax: +61 2 9871 6941. E-mail address:
[email protected] (F.A. Ximenes).
0956-053X/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.wasman.2007.11.006
2002). Methane is a greenhouse gas, with a global warming potential 23 times higher than carbon dioxide. It has been estimated that methane emissions from landfills may account for up to 20% of global anthropogenic methane emissions (IPCC, 2002). It is not certain to what extent decomposition of wood products contributes to methane emissions from landfills. Depending on their rate of decomposition, wood products placed in a landfill may represent either a carbon sink or a source of greenhouse gases. Knowledge about their rate and extent of decomposition is important for several reasons, including: (1) development of more accurate national greenhouse gas inventories; (2) accurate assessment of the contribution of wood and paper products to long-term sequestration of carbon; (3) calibration of models used to predict recoverable methane and (4) development of waste diversion policies. Most field studies on the anaerobic decomposition of organic materials in landfills have focused on the refuse as a whole, rather than individual components of the waste stream. Therefore, there are very few field data specifically related to wood products. The estimates of the decomposition of wood products in landfills used in national greenhouse inventories are typically based on theoretical
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stoichiometric calculations or laboratory studies (US EPA, 2005; IPCC, 2006). Barlaz et al. (1990) suggested that stoichiometric calculations may overestimate methane generation by up to 50%. Bogner (1992) observed that methane generation data from laboratory-based experiments under optimised conditions often overestimate actual methane production in landfills. The intergovernmental panel on climate change (IPCC) method for calculation of the extent of decomposition of wood in a landfill under anaerobic conditions utilises two factors: the degradable organic carbon (DOC), that is, the fraction of wood that may be subjected to biochemical decomposition, and the fraction of DOC that is dissimilated (DOCf), that is, the fraction of DOC that is actually decomposed in a landfill. The default DOC and DOCf values recommended by IPCC (2006) for dry wood in a landfill are both 0.5. These values have been adopted in Australia’s National Greenhouse Accounts for the Waste sector (AGO, 2006). Together, the two factors lead to an assumed loss of 50% of the carbon originally in the wood. In contrast, Micales and Skog (1997) conducted an extensive literature review and, based primarily on estimations of methane generation from commercial models, concluded that only 0–3% of the carbon from wood products buried in a landfill is released into the atmosphere. However, these estimates have not been confirmed from field-based observations. The aim of this study was to recover wood products from landfills and analyse them chemically to provide field-based data on the decomposition of wood products in Australian landfills. This paper reports on the findings of excavations and chemical analyses of the wood products recovered from three sites at two landfill facilities in Sydney. The widely divergent estimates of decomposition of wood products in landfills reflect the very limited understanding of the nature of their degradation in landfills. The long-term effects of burial in landfills will not be known for many decades as the advent of managed landfills is relatively recent. However, experiments conducted under anaerobic conditions and studies of degraded archaeological wood provide useful insights into the potential mechanisms involved. Therefore it is informative to review briefly the knowledge and methodology developed through observations of wood decomposition in anaerobic environments. 1.1. Background on decomposition of wood under anaerobic conditions Fungal decay of wood and wood products only occurs in the presence of oxygen (Blanchette et al., 1990); therefore, any biological degradation of the wood and paper samples during burial in anaerobic landfills occurs as a result of bacterial activity, not fungal. Bacteria are known to degrade wood at a slower rate compared to fungi (Blanchette et al., 1990), and there is very little understand-
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ing of the mechanisms of bacterial attack of wood under anaerobic conditions in landfills. Blanchette (2000) described three main groups of bacteria that can degrade wood (primarily the secondary wood cell wall layers): erosion, cavitation and tunnelling bacteria. Erosion bacteria deplete cellulose and hemicellulose from the wood cell wall, producing small troughs. Tunnelling bacteria produce minute tunnels primarily within the secondary cell wall, whereas cavitation bacteria form small diamond-shaped cavities perpendicular to the long direction of the fibre. Tunnelling type bacteria have been found to be concentrated mostly on surface layers of waterlogged wood and not in the inner wood where oxygen is limited (Bjo¨rdal et al., 1999). In oxygen-limited environments, erosion bacteria are reported to be the most common type of bacterial degrader (Blanchette, 2000). 1.2. Degradation of wood under anaerobic conditions in laboratory experiments A few studies have described the anaerobic decomposition of wood products under controlled conditions in the laboratory. Peltola et al. (2000) examined the anaerobic degradability of plywood (plain and overlaid with phenolic surface film) and chipboard at 33 °C using municipal digested sludge as the seed. The samples were ground to dust and less than 1 mg of substrate was tested. The combined methane and carbon dioxide yields (they were emitted in approximately equal proportions) were 55% of the carbon content of plain plywood and 33% of the carbon content of surface film-coated plywood. The combined methane and carbon dioxide yield for chipboard was approximately 5%. The authors suggested that the high anaerobic mineralization of plywood was probably due to the perforation of the lignified cell walls due to mechanical grinding (although this does not explain the low yields for chipboard, as it was also subjected to mechanical grinding). Such effect would obviously not occur in plywood that is placed in landfills. Differences in wood species used in the manufacture of both materials (the species composition of the chipboard was not mentioned) and type of adhesive used may help explain the differences in decomposition. Tong et al. (1990) subjected finely ground white fir (Abies concolor Lindl.) to a biochemical methane potential test (BMP) conducted at 35 °C. The maximum methane conversion efficiency (methane produced divided by the theoretical maximum methane potential) for white fir was less than 10%. This suggests that a significant amount of the cellulose and hemicellulose was not available for anaerobic bacterial decomposition, even under optimised testing conditions. The authors suggested that lignin may limit the availability of carbohydrates by forming a barrier to enzymes or by forming chemical bonds with the carbohydrates that are not susceptible to hydrolysis. The results of Peltola et al. (2000) and Tong et al. (1990) suggest that even after increasing the surface area of the wood (ground wood has a greater surface area than solid
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wood) exposed to optimal conditions for degradation to occur, a significant proportion of the carbon does not degrade. Holt and Jones (1983) buried beech (Fagus sylvatica) and Scots pine (Pinus sylvestris) blocks in anaerobic muds for up to 18 months and recorded weight losses ranging from 0 to a maximum of 4.4%, with extensive superficial erosion of the cell walls by rod-shaped bacteria. The estimation of greenhouse gas emissions from wood and paper products in landfills in the US (US EPA, 1998) is based on the results of laboratory experiments in which paper products and tree branches were ground and degraded anaerobically in 2-l jars (Barlaz et al., 1997). The test conditions were designed to measure the maximum methane production of each component. The extent of decomposition (calculated as the measured methane yield divided by the yield calculated assuming conversion of 100% of the cellulose and hemicellulose to methane and carbon dioxide) of branches, old newspaper and office paper was 27.8%, 31.1% and 54.6%, respectively. The results for the paper types used in the study were applied by the US EPA to other types of paper products, and the results for branches were used as proxy for solid wood and medium density fibreboard (MDF). This does not seem appropriate since the structure and composition of solid wood and panel products is inherently different from that of tree branches, and different types of paper products vary considerably in their composition.
mately 100 m. Although no significant decay was found in spruce, the analysis of the buried red alder and white oak revealed extensive decay, with the hemicellulose being preferentially decayed. Sola´r et al. (1987) also found preferential decomposition of hemicellulose in 8138-year old subfossile oak (Quercus robur) excavated from 16 m below the ground. Boutelje and Go¨ransson (1975) found evidence of selective degradation of cellulose of Scots pine (Pinus sylvestris) buried in semi-anaerobic conditions. It is important to note that degradation patterns may vary according to the environment (e.g., aquatic, mudlands, landfills), as the bacteria type and surrounding matrix may also vary. One potential difficulty in interpreting analyses of archaeological woods (as well as more recently buried wood) is that one cannot rule out the possibility of aerobic degradation prior to onset of anaerobic conditions. The distinct fungal degradation patterns, which would indicate aerobic degradation, may not be obvious after a very long period of time.
1.3. Degradation in archaeological wood
The Lucas heights landfill facility is located in the southern suburbs of Sydney, Australia. The LH sites were selected based on the landfill operators’ knowledge of historical waste deposition. Two sites (A and B) at the LH landfill facility were excavated with a bucket excavator and sampled in May 2001. Site A contained material that was deposited in 1982 and had been buried for 19 years at the time of excavation. The material had been compacted on deposition. Methane was being collected from this site and used to generate electricity at the time of excavation. The site was excavated to the base, 6 m below the surface of the cover soil. Material was segregated in 0.5 m depth increments. Site B contained material that was deposited in 1972 and had been buried for 29 years at the time of excavation. The refuse had not been compacted on deposition. The site contained a recirculating leachate system. The site was excavated to the base at 4 m below the surface of the cover soil, and the excavated contents were segregated in 0.5 m depth increments.
Archaeological wood can be defined as dead wood, that may or may not have been modified for or by use, and that was discarded by intent or accident into a specific natural environment (Florian, 1990). Several papers have reported degradation of archaeological wood found in anaerobic or semi-anaerobic conditions, such as waterlogged environments (e.g., shipwrecks) and intertidal sites. It has commonly been reported that bacterial attack is responsible for the deterioration of the archaeological wood carbohydrates, leaving a structure consisting mainly of residual lignin (Blanchette, 2000). Depth is an important factor often correlated with the severity of decay in archaeological samples. Boutelje and Go¨ransson (1975) sampled 70-year old foundation wooden piles made of spruce 4–8 m below the groundwater table in Gothenburg, Sweden. The authors did not detect any decay in those samples, whereas samples of piles taken closer to the surface (0.4 m) showed initial signs of decay. Based on the analysis of discs from waterlogged 1200-year old softwood poles, Bjo¨rdal et al. (2000) suggest that even small differences in depth of burial of wood may result in significant variations in decay. The pole samples analysed by Bjo¨rdal et al. were not subjected to completely anaerobic conditions. Hedges et al. (1985) analysed the decomposition of 25,000-year old buried spruce (Picea sitchensis), red alder (Alnus rubra) and white oak (Quercus spp.) recovered at a sediment depth of approxi-
2. Methods The landfill sites contained newspapers, magazines, calendars and telephone books that were used to establish the year of disposal of the wood products recovered from each depth. 2.1. Lucas heights (LH)
2.2. Sydney park (SP) The Sydney park landfill is located in central Sydney. The SP site was selected based on old aerial photographs of the site when it was still active. The site had ceased to receive waste 46 years prior to sampling. The refuse had not been compacted on deposition. The site was excavated
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with a bucket excavator in December 2002. After opening a hole through the concrete cover, the site was excavated to the base, approximately 6 m from the surface. The excavated contents were segregated in 0.5 m depth increments. Only a small quantity of paper products was recovered from SP, most likely because it was common at the time that the site was active to burn organic materials, particularly paper, in incinerators or for home heating. The newspaper samples recovered were used to establish the year of disposal of the wood products. 2.3. Physical and chemical analyses The redox potential of the matrix at SP was measured using an Ag/AgCl reference electrode and platinum electrodes (the equipment was not used for the LH excavations). Measured potentials were converted to soil Eh by adding 199 mV to account for the difference between the Ag/AgCl electrode and the standard hydrogen electrode. The moisture content of twenty wood samples from LH (10 from Site A and 10 from Site B) and 48 samples from SP (randomly selected) was determined. The samples were sealed in plastic bags immediately after removal from the landfill. The moisture content of the samples was determined by oven-drying at 103 ± 2 °C until constant weight in accordance with AS/NZS 1080.1 (1997). The moisture content of the samples was expressed as moisture content ð%Þ ¼ ððfresh weightðgÞ oven-dry weightðgÞÞ =fresh weightðgÞÞ 100 The LH and SP wood samples were identified by a wood anatomist. The samples were grouped initially into hardwoods and softwoods. Where possible they were further identified to species using standard techniques based on light microscopy, colour and density. Colour is a key factor in identification, and dark staining of some of the hardwood samples (particularly from the Eucalyptus genus) prevented them from being identified to species. In that case, samples were grouped according to their basic density (medium density = 600–900 kg/m3 and high density > 900 kg/m3). A total of 77 landfill samples was identified to species, as outlined in Table 1, and analysed for chemical composition. A total of 84 matched control samples of the same or similar species to those recovered from the landfill sites, obtained from the Forests NSW wood collection (collected from trees grown early in the 20th century), was also analysed for chemical composition (Table 1). It is reasonable to assume that most of the wood products recovered from the landfills included in this study were manufactured from trees harvested early in the 20th century or even earlier. Thus, there is a strong possibility that the control wood samples and the landfill wood samples were sourced from similar stocks. Because treated wood products have been used for residential purposes in Australia only since the 1950s, it was
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Table 1 Number of samples identified to species and analysed for chemical composition Wood type
Lucas heights
Sydney park
Matched controlsa
Site A
Site B
Hardwoods Softwoods
10 10
6 11
25 16
45 39
Total
20
17
40
84
a
Samples of the same or similar species to those recovered from the landfill sites, obtained from the Forests NSW wood collection (collected from trees grown early in the 20th century). The number of control samples used for the analyses was adjusted so that they replicated the relative proportion of the hardwood and softwood species recovered from landfill.
considered unlikely, though possible, that treated wood products had been disposed at the landfill sites studied. Tests for As, Cu, and Cr were conducted on 25 samples that had tested positive with a chrome azul method, and no significant levels of these metals were found. Prior to chemical analysis, the landfill and control samples were air-dried and then oven-dried at 40 °C before being milled to pass a 0.75 mm screen. Carbon concentration was determined by combustion decomposition followed by infra red detection using a LECO SC444 carbon analyser. The following methods were used to determine the cellulose, hemicellulose and lignin fractions: Enzymatic neutral detergent fibre (ENDF): to determine the sum of cellulose, hemicellulose and lignin (McQueen and Nicholson, 1979). Acid detergent fibre (ADF): to determine the sum of cellulose and lignin fraction in plant material (AOAC, 1995). The procedure for detergent fibre analysis is similar to crude fibre analysis as it measures the residue of plant cells after fractionation using detergent solutions to solubilise protein and starch. Klasson lignin: (AOAC, 1995). Klason lignin (or 72% (v/v) H2SO4 acid procedure) is the most common method utilised to determine lignin. This method dissolves away all of the polysaccharides, leaving lignin as an insoluble residue (Browning, 1967). The total weight of carbon in each sample was calculated assuming that the carbon concentration of cellulose and hemicellulose were, respectively, 44.4% (based on the molecular weight of carbon in the cellulose unit) and 45% (based on the molecular weight of carbon in pentosan and hexosan units). The carbon concentration of lignin was assumed to be 63% for softwoods and 58% for hardwoods (Fengel and Wegener, 1984). If the difference between the mean lignin concentration of the landfill samples and that of the control samples
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was not statistically significant (P > 0.05), it was assumed that no significant loss of mass had occurred. If the differences were significant, then the loss of carbon mass was calculated as follows (Hedges et al., 1985):
assessed using normal probability plots and histograms prior to testing (no transformation was required).
(1) Because lignin is considered recalcitrant under anaerobic conditions (e.g., Barlaz et al., 1990; Ham et al., 1993; Wang et al., 1994), a measure of the loss of wood mass is given by the factor (F) by which the lignin concentration is increased in the landfill samples: F = lignin % (control)/lignin % (landfill). (2) Proportion of cellulose and hemicellulose that is preserved in landfill (P1 and P2): P1 = (cellulose % (landfill)/cellulose % (control)) F. P2 = (hemicellulose % (landfill)/hemicellulose % (control)) F. A decrease in ‘‘P1” and ‘‘P2” is followed by a decrease in ‘‘F”. (3) Total carbon loss (C loss): C loss (% of original sample mass) = [cellulose % (control) carbon fraction in cellulose (1 P1)] + [hemicellulose % (control) carbon fraction in hemicellulose (1 P2)].
All wood samples recovered were extremely wet (above their fibre-saturation point of 30%), with moisture content ranging from 42.5% to 62.5% at LH and 41.6% to 66.8% at SP (Tables 2 and 3). The mean moisture content of the softwoods was significantly higher than that of the hardwoods for both sites. No trend in the moisture content of the wood with depth was observed (Table 3). Moisture is acknowledged as a critical factor influencing decomposition rates in landfills (Bogner and Spokas, 1993). As the refuse moisture content is increased, the opportunity for contact between microorganisms and their substrates is also increased, which should enhance microbial activity (Barlaz et al., 1990). Although not quantified, the refuse as a whole in all of the excavations was typically either very damp or dripping wet (Fig. 1a and b). However, other factors such as compaction, refuse age and refuse composition influence the optimal moisture content for decomposition to occur (Barlaz et al., 1990). Shearer (2001), based on lysimeter studies, suggests a minimum moisture content for degradation of 25%, and that 40–70% moisture content is optimal for degradation. The average moisture content of the wood samples recovered from LH and SP landfills and the standard deviation suggests that lack of moisture was not a limiting factor for decomposition to occur (Tables 2 and 3). The redox potential measured at SP ranged from 640 mV to 40 mV, corresponding to strongly reducing conditions. The low redox potential at SP and the fact that methane was being collected at LH indicate that both landfill sites were anaerobic. The majority of the wood samples recovered after 19–46 years in the landfills showed no obvious visual signs of deterioration (Fig. 2). Rough sawn and dressed boards, mouldings, furniture components and other articles, including broom and tool handles, were readily identifiable among the wood products removed from the landfill sites. A variety of paper products including newspapers, magazines, telephone books, cardboard and office paper were also recovered from Lucas heights (Fig. 3).
2.4. Statistical analyses All statistical analyses were performed using SAS 8.02, and a one-way t-test was used to compare the means, with a significance level of 0.05. The choice of one-sample t-test was based on the fact that the research question asked was: ‘‘are landfill samples more degraded than control samples?” i.e., not whether they were ‘‘less” degraded than control samples as well. In all comparisons, the control samples were selected to match the species composition in the landfill samples. It was assumed that the control group was the population mean and that it contained no error. Initially the analyses were carried out on a species basis, but in view of the low sample size and high variability, the samples were grouped as softwoods and hardwoods for the analyses. The number of control samples used for the analyses was adjusted so that they replicated the relative proportion of the hardwood and softwood species recovered from landfill. The assumption of normality was
3. Results and discussion
Table 2 Moisture content of LH wood samples according to wood type Site
Statistics
Softwoods
Hardwoods
Softwoods and hardwoods
Site A (2.0–2.5 m)
Moisture content (%)
Mean N SD
51 3 1.8
42.5 7 5.4
45.1 10 6.1
Site B (2.0–3.0 m)
Moisture content (%)
Mean N SD
62.5 6 8.2
53.2 4 6.7
58.8 10 8.7
Total
Moisture content (%)
Mean N SD
58.7 9 8.7
46.4 11 7.8
51.9 20 10.1
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Table 3 Moisture content of SP wood samples according to wood type and depth Depth (m)
Statistics
Softwoods
Hardwoods
Softwoods and hardwoods
2.5
Moisture content (%)
Mean N SD
64.3 3 4.6
49.2 5 13.1
54.8 8 12.9
3.5
Moisture content (%)
Mean N SD
59.5 2 3.0
41.9 9 5.8
45.1 11 8.9
4.5
Moisture content (%)
Mean N SD
66.3 4 7.2
48.4 6 9.7
55.6 10 12.5
5.5
Moisture content (%)
Mean N SD
67.4 7 8.0
41.6 7 3.1
54.5 14 14.6
6.5
Moisture content (%)
Mean N SD
66.8 5 9.7
– – –
66.8 5 9.7
Total
Moisture content (%)
Mean N SD
65.2 21 7.4
43.5 27 7.7
53.9 48 13.2
Fig. 1. (a) Typical waste excavated from Lucas heights and (b) typical waste excavated from Sydney park.
Fig. 2. Wood samples from Sydney park.
The wood samples were subsequently identified to include a range of Australian and exotic hardwood and softwood species groups (Table 4). The samples recovered from the landfills included species that are known, from studies of performance in service (Bootle, 1983), to be very durable, such as ironbark, and those known to have low durability, such as those in the radiata pine group (Table 4). There was no visual evidence that greater deterioration had occurred in the less durable species recovered from LH and SP. Differences in the degree of deterioration would be expected between less durable and more durable species under aerobic conditions, as seen for example in brown-rotted wood (brown-rot fungi selectively degrade cellulose and hemicellulose). If significant bacterial degradation of wood occurs in landfills under anaerobic condi-
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recovered in both studies were not chemically analysed to determine depletion of specific components. Kinman et al. (1990) noted that the biological degradation of the carbon containing materials (including wood) after 17–19 years burial had occurred at a relatively slow rate. Bogner and Spokas (1993) suggested that more than 75% of all of the carbon deposited in landfills remains in sedimentary storage. Degradation is a result of changes in the chemical structure, which does not necessarily result in physical mass loss, and therefore may not necessarily equate to loss of carbon. Thus, there may have been greater degradation, in terms of change to chemical structure, than is indicated by the loss of carbon. On the other hand, the loss of carbon indicated by the decline in cellulose and hemicellulose content may overestimate the carbon lost from the landfill; it is possible that the breakdown products of cellulose and hemicellulose are in fact present in the matrix surrounding the wood rather than being released as carbon dioxide or methane. Degradation may also be of a chemical nature in landfills, in which case, no carbon is released as carbon dioxide or methane either. This is particularly relevant for hemicelluloses, which have less-ordered structures, higher solubilities and are more readily hydrolysed than celluloses (Hedges, 1990). For the purposes of this paper, it was assumed that all physical mass loss resulted in release of carbon from the landfill. The carbon, cellulose, hemicellulose and lignin concentrations of the control samples and the samples collected from LH Sites A and B and SP are summarised in Tables 5 and 6. The carbon concentration of the landfill samples was not significantly different from that of control samples (Tables 5 and 6). It is generally accepted that lignin polymers in wood products do not decompose at all in landfills. Since cellulose and hemicellulose account for the bulk of the remaining mass of wood, a significant increase in lignin concentrations must result almost exclusively from selective polysaccharide (cellulose and hemicellulose) removal (Hedges, 1990). As the lignin concentration of the control samples was not significantly different from that of landfill
Fig. 3. Paper samples from Lucas heights.
tions (and the lignin is preserved), one would expect to find wood products in landfills that would resemble aerobically degraded, brown-rotted wood. No such samples were observed in the three excavations reported here. It is not impossible that some wood product samples may have been degraded beyond recognition. However, in a biologically active system, one would not expect to find only extremes in terms of degradation (samples that have not degraded at all or only to a small extent, and samples that have been completely degraded) without finding the middle ground, i.e., samples that have been degraded extensively but not beyond recognition. Despite evidence that both landfills accepted typical mixed municipal solid waste during their operational lives, no identifiable food scraps were found, which is a strong indication that both landfills were biologically active for at least some time. The state of the contents recovered in landfill excavations has been reported in some previous studies. Walsh and LaFleur (1995) recovered sound wood samples buried for up to 100 years from excavations in a New York landfill. Wood accounted for 12.3% of the refuse sampled after 17–19 years burial in the Mallard North Landfill in Chicago (Kinman et al., 1990). However, the wood samples Table 4 Identification of wood from Lucas heights Sites A and B and Sydney park Site
Wood type
Site
Wood type
Softwood LHA X X X X X X X
LHB
SP
X
X X X X X
X X X
X X a
Botanical name Agathis sp. Callitris sp. Larix sp. Picea sp. Pinus sp. Pseudotsuga sp. Tsuga sp. Pinus sp. Pinus sp. Pinus sp. Pinus sp.
Hardwood Common name Kauri Cypress pine Larch Spruce Sylvestris group Douglas fir Hemlock Radiata pine group White pine group Ponderosa group Taeda group
LHA
LHB
X X
SP
Eucalyptus sp. medium density
X X X X
Euc. high density Eucalyptus sp.a Eucalyptus sp.a Lophostemon sp. Shorea sp. Gonystylus sp. Dipterocarpus sp. Doryphora sp.
X X X
Common name
a
X
X
Samples could only be identified to genus due to staining of the samples in the landfill.
Botanical name
Ironbark group Stringybark group Brush box Meranti Ramin Keruing Sassafras
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Table 5 Summary of analytical data for control samples and landfill samples from LH Sites A and B (mean and SD) Component
Statistics
LH A softwoods a
LH A hardwoods a
LH B softwoods
Control N = 22
Landfill N = 10
Control N = 11
Landfill N = 10
Controla N = 18
Landfill N = 11
Carbon (% oven-dry)
Mean SD CV (%)
50.9 1.5 2.9
50.7 3.3 6.6
49.9 1.2 2.4
49.6 1.4 2.7
50.8 1.2 2.4
50.8 1.2 2.4
Cellulose (% oven-dry)
Mean SD CV (%)
46.6 4.6 9.9
51.1 6.7 13.0
55.7 5.1 9.2
53.9 6.0 11.1
49.4 3.1 6.3
52.6 3.1 6.0
Hemicellulose (% oven-dry)
Mean SD CV (%)
12.8 2.7 21.1
8.7 2.2 25.3
15.9 5.1 32.1
9.3 3.4 36.5
13.4 2.9 2.6
11.8 2.4 20.6
Lignin (% oven-dry)
Mean SD CV (%)
31.6 3.4 10.8
29.6 4.3 14.4
23.3 7.8 33.5
26.5 5.7 21.4
30.0 3.3 11.0
30.7 3.2 10.4
pb
0.178
0.108
0.48
a
The number of control samples used in the analyses was artificially adjusted to match the proportion of each species found in landfill, and thus minimise bias in the chemical analyses. b Significance of the difference between the mean lignin concentration of the landfill samples and that of the control samples.
Table 6 Summary of analytical data for control and landfill samples (combined LH Sites A and B and Sydney park), (mean and SD) Component
LH total softwoods a
LH total hardwoods a
SP softwoods a
SP hardwoods
Control N = 22
Landfill N = 10
Control N = 11
Landfill N = 10
Control N = 18
Landfill N = 11
Controla N = 24
Landfill N = 25
Carbon (% oven-dry)
Mean SD CV (%)
50.8 1.4 2.8
50.8 2.4 4.7
50.3 1.5 3.0
50.3 1.6 3.1
50.2 0.8 1.6
50.8 1.1 2.1
51.3 0.9 1.7
50.8 1.3 2.6
Cellulose (% oven-dry)
Mean SD CV (%)
48.0 4.6 9.6
51.9 5.1 9.7
55.7 6.2 11.1
55.8 5.7 10.1
51.2 3.0 5.9
46.8 3.9 8.3
52.4 4.7 9.0
49.1 6.5 13.0
Hemi cellulose (% oven-dry)
Mean SD CV (%)
13.1 2.8 21.4
10.3 2.8 26.7
15.3 3.9 25.5
9.4 4.3 46.1
13.7 2.7 19.7
10.5 1.8 17.2
15.6 3.9 25.0
14.4 2.8 19.3
Lignin (% oven-dry)
Mean SD CV (%) pb
30.8 3.7 12.0 0.49
30.2 3.7 12.2
23.8 7.8 32.8 0.27
25.5 5.9 23.0
29.6 1.8 6.1 <0.0001
37.1 2.9 7.9
23.8 4.6 19.3 0.0005
29.5 6.9 23.1
a The number of control samples used in the analyses was artificially adjusted to match the proportion of each species found in landfill, and thus minimise bias in the chemical analyses. b Significance of the difference between the mean lignin concentration of the landfill samples and that of the control samples.
samples from Lucas heights (Table 5), no significant loss of mass, and therefore, of carbon was assumed to have occurred in those samples. There were insufficient hardwood samples recovered from LH B to allow meaningful comparisons between landfill and control samples. The lignin concentration of the SP samples was significantly higher than that of the control samples (Table 6). The coefficient of variation of the chemical composition of the landfill samples was typically similar to that of the corresponding control samples, particularly for lignin (Tables 5 and 6). The calculated loss of carbon as a percentage of dry wood mass for SP softwoods and hardwoods was 9.1% and 8.7%, respectively (Table 7), equating to 18%
and 17% of their original carbon content, respectively. Published estimates of decomposition of carbon in wood products in landfills range from 0–3% (Micales and Skog, 1997, based on published estimates of methane yields), to 35% (Mann and Spath, 2001, based on the assumption that 50% of the cellulose and hemicellulose fractions of wood degraded in landfills), to 56% (Bingemer and Crutzen, 1987, based on stoichiometric calculations) of the total carbon in the wood products. The IPCC (2006) assumes that 50% of the carbon originally in the wood decomposes in landfills. Several factors may contribute to the measured differences in degradation of wood products from LH and SP.
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Table 7 Lignin changes, proportion of cellulose and hemicellulose (P1 and P2) preserved and carbon losses for wood samples from Lucas heights (LH) and Sydney park (SP) landfills Component
LH total softwoods
LH total hardwoods
SP softwoods
SP hardwoods
F P1 (%) P2 (%) Carbon loss (% oven-dry wood mass)
1.02 100 100 –
0.93 100a 100a –
0.80 72.9 61.5 9.1
0.81 75.7 74.4 8.7
a
As the change in lignin concentration (F) was not statistically significant, P1 and P2 were considered to be 100%.
Those include but are not restricted to: (1) the longer burial time at SP; (2) differences in the original waste mix; (3) potential differences in pH (not measured for SP) and temperature; (4) potential loss of C due to fungal degradation prior to burial at SP; and (5) anaerobic nature of the site (waste at SP may have been exposed to oxygen for some time before the conditions became anaerobic, allowing fungal activity to occur). The number of samples recovered at each depth was insufficient to allow meaningful analyses of the effect of depth on degradation. However, the low variability of the chemical composition data between individual samples from the same species collected at different depths (data not shown), as well as their sound appearance, suggested that depth did not significantly influence the level of degradation. For SP hardwood samples, the proportional loss of cellulose, derived according to Section 2, was practically the same as that of SP hardwood hemicellulose (approximately 25%). The proportional loss of hemicellulose from the SP softwood samples (37.7%) was higher than that of the SP hardwood cellulose (27%). This corresponds with the observation that, in general, bacteria degrade hemicellulose faster than cellulose (Fengel and Wegener, 1984). Hoffmann (1981) analysed the carbohydrate fractions of samples of archaeological European oak up to 2000 years old and found that both cellulose and hemicelluloses had been removed at about the same ratio as they were assumed to be present initially in wood. There are other reports, however, of preferential degradation of cellulose in submerged pine pilings (Boutelje and Go¨ransson, 1975) and of hemicellulose in buried hardwoods (Kohara, 1956; Hedges et al., 1985). Studies describing the relative loss of cellulose and hemicellulose for wood buried in landfills were not found in the literature. The low decomposition of wood products reported here may be primarily due to the fact that anaerobic bacteria are not able to reach portions of the cellulose and hemicellulose that are encrusted in lignin (Ham et al., 1993; Wang et al., 1994). Schmidt (1980) conducted an experiment in which wood samples with original and reduced lignin content were subjected to bacterial attack both under aerobic and anaerobic conditions. Those samples with the original
lignin content were not degraded, whereas those samples with reduced lignin content had considerable weight loss. The experiments suggested that even a small reduction in the lignin content of the samples may be enough to result in considerable decay. The degree to which the carbohydrates in wood are enveloped by lignin may be a key factor determining potential differences in loss of carbon from different wood types. 3.1. Potential implications of results The carbon losses from Sydney park samples (Table 7) translate into a maximum DOCf for wood products of 0.3, assuming a DOC of 0.3 (if lignin is not considered to be ‘‘degradable”), or a DOCf of 0.18, assuming a DOC of 0.5 (if lignin is considered to be ‘‘degradable”). Barlaz (2004, 2006) listed published DOC values for wood products ranging from 0.18 to 0.3 (excluding lignin) and DOCf values ranging from 0.5 to 0.6. The DOCf value for wood recovered from Sydney park samples is 2.8 times lower than that adopted by the IPCC (2006). The IPCC allows the use of country-specific DOCf values rather than the suggested default values, provided they are based on welldocumented research (IPCC, 2006). It is important to note that there is currently insufficient field evidence to determine what proportion of carbon from wood products potentially lost through degradation is emitted as carbon dioxide and as methane. The true greenhouse impact of wood products in landfill will depend on how much carbon is released as methane, and also whether any methane generated is simply released into the atmosphere, flared or utilised to generate electricity. If the low decomposition values suggested here are confirmed by further excavations, the implications for reporting of national greenhouse gas emissions will be significant to those countries that rely on landfills as a major disposal option for redundant wood products. These decomposition data should also assist in achieving acknowledgement of the role that wood products play in extending the benefits of carbon sequestration in trees, since potentially a large carbon reservoir may be retained in landfills. Skog and Nicholson (2000) estimate that the projected annual rate of accumulation of carbon in wood and paper products in landfills and in products in service in the USA (75 million tonnes in 2040) is due entirely to the increasing rate of accumulation in landfills. The low decomposition of wood products in landfill may help to explain the fact that methane yields from landfill are often lower than predicted (Barlaz et al., 1990; Augenstein and Pacey, 1991; Bogner, 1992). Bogner (1992) reported rates of biogas production from refuse in landfills ranging from 0.007 m3 (dry) kg1 year1 to more than 1.0 m3 (dry) kg1 year1, with the higher rates resulting from laboratory studies and the lower rates from field pumping tests. There are several differences between methane production from refuse placed in anaerobic digesters and in landfills; refuse in landfills have large particle size,
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generally low water content and poor mixing, all resulting in low methane production rates (Bogner, 1992). The discrepancies in the methane extracted from different landfills may be due to local variations in the amount of food scrap, garden waste and possibly paper products placed in landfills, as those materials are potentially more readily degradable than wood products. Combined, those materials account for approximately 80% of the mass of organic materials placed in landfills in Australia (AGO, 2006). In Australia, the Commonwealth Government has established a scheme that places a legal liability on wholesale purchasers of electricity to proportionately contribute towards the generation of renewable energy (Mandatory Renewable Energy Targets). This scheme is one of the major drivers for installation of facilities to generate electricity from the methane produced by landfills (AGO, 2004). If waste is diverted from landfill on the basis of unrealistically high methane emission factors for wood and paper products, it may create incentives for the uptake of alternative practices (such as incineration), in effect resulting in net increases in greenhouse gas emissions. 3.2. Recommendations for future research This research was limited to three sites at two landfills with different management practices, and the chemical analyses were restricted to solid wood products. More excavations are planned to further define the influence of landfill age, type, environment and management practices on the extent of carbon loss from both solid wood products and a wide range of paper types buried in landfills in Australia. The principles and methodologies used to guide research on the decomposition of forest products in landfills were the topic of a recent workshop (Ximenes et al., 2005a,b). Major areas for further work include but are not restricted to (1) effect of the sapwood and heartwood content of the solid wood samples on their chemical composition; (2) the use of microscopy to bring a better understanding of the mechanisms of degradation, its nature (whether fungal or bacterial, based on decay patterns) and extent; (3) chemical analyses of samples of fresh materials degraded anaerobically under optimal conditions in the laboratory; (4) use of nuclear magnetic resonance (NMR) and tetramethylammonium hydroxide thermochemolysis (TMAH) in the chemical characterization of degraded samples from old landfills; and (5) nature of the degradation of hemicellulose, i.e., whether it is of a microbial or chemical nature. 4. Conclusions The methodology used in this paper allowed, for the first time, the determination of the decomposition of wood products recovered from landfills. Examination of a wide range of wood products recovered after 19 and 29 years burial in two sites at the Lucas heights landfill facility found no significant loss of dry mass. However, at Sydney
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park, where refuse had been buried for 46 years, the loss of carbon (as a percentage of dry biomass) was 8.7% for hardwoods and 9.1% for softwoods, corresponding to 18% and 17% of their original carbon content, respectively. These results support previously suggested evidence of slow decomposition of wood in landfills. Acknowledgements This research was funded by the CRC for Greenhouse Accounting, Forests NSW and the Chemistry Centre of Western Australia. Access to the landfill sites and excavation equipment was provided by Waste Services New South Wales and South Sydney City Council. The planning and execution of this research was greatly assisted by the guidance and advice offered by the members of the Landfill Research Steering Committee. The authors would like to acknowledge the assistance provided by Dr. Andrew Haywood (formerly Forests NSW) with the statistical analyses; Mark Stevens, from the City of Sydney Archives, in the identification of the Sydney park site and Lyndell Dunne, from the State Library of New South Wales, for the identification of the newspaper that helped dating the samples recovered from Sydney park. Helpful comments on this manuscript provided by Dr. Morton Barlaz (North Carolina State University), and by anonymous reviewers are gratefully acknowledged. References AGO (Australian Greenhouse Office), 2004. Waste Sector – Greenhouse Gas Emissions Projections 2004. Commonwealth of Australia, 43 p. AGO (Australian Greenhouse Office), 2006. Australian National Inventory Report 2004, vol. 2. The Australian Government Submission to the UN Framework Convention on Climate Change, Commonwealth of Australia, p. 166–173.
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