Impact assessment of intermediate soil cover on landfill stabilization by characterizing landfilled municipal solid waste

Journal of Environmental Management 128 (2013) 259e265

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Impact assessment of intermediate soil cover on landfill stabilization by characterizing landfilled municipal solid waste Guangxia Qi a, b, Dongbei Yue a, b, *, Jianguo Liu a, b, Rui Li a, b, Xiaochong Shi c, Liang He d, Jingting Guo d, Haomei Miao d, Yongfeng Nie a, b a

School of Environment, Tsinghua University, Beijing 100084, China Key Laboratory for Solid Waste Management and Environment Safety, Ministry of Education of China, Tsinghua University, Beijing 100084, China College of Marine Life Sciences, Ocean University of China, Qingdao 266003, China d BESG Environmental Engineering Co., Ltd. Beijing 100101, China b c

a r t i c l e i n f o

a b s t r a c t

Article history: Received 11 December 2012 Received in revised form 4 May 2013 Accepted 11 May 2013 Available online

Waste samples at different depths of a covered municipal solid waste (MSW) landfill in Beijing, China, were excavated and characterized to investigate the impact of intermediate soil cover on waste stabilization. A comparatively high amount of unstable organic matter with 83.3 g kg
1 dry weight (dw) total organic carbon was detected in the 6-year-old MSW, where toxic inorganic elements containing As, Cd, Cr, Cu, Mn, Ni, Pb, and Zn of 10.1, 0.98, 85.49, 259.7, 530.4, 30.5, 84.0, and 981.7 mg kg
1 dw, respectively, largely accumulated because of the barrier effect of intermediate soil cover. This accumulation resulted in decreased microbial activities. The intermediate soil cover also caused significant reduction in moisture in MSW under the soil layer, which was as low as 25.9%, and led to inefficient biodegradation of 8- and 10-year-old MSW. Therefore, intermediate soil cover with low permeability seems to act as a barrier that divides a landfill into two landfill cells with different degradation processes by restraining water flow and hazardous matter. Ó 2013 Elsevier Ltd. All rights reserved.

Keywords: Municipal solid waste Landfill Intermediate cover Stabilization

1. Introduction Over 80% of municipal solid waste (MSW) in China is still being disposed of in anaerobic landfills at present (He et al., 2011; Hong et al., 2010). Moisture is essential for metabolism of all microorganisms (Bäumler and Kögel-Knabaner, 2008; Bilgili et al., 2007; Staub et al., 2010; Valencia et al., 2009) and is accordingly acknowledged as one of the most important factors influencing MSW degradation inside a landfill site. Relatively high moisture content improves the mixing and general availability of nutrients and carbon-rich organic matter, and thus stimulates bacterial growth directly and leads to enhanced degradation of waste materials (Wreford et al., 2000). At the same time, however, moisture can dissolve and transport organic and inorganic compounds and some metabolic inhibitors (Rees, 1980), such as heavy metals. Water may infiltrate through solid waste as an unsaturated flow inside a landfill (Korfiatis et al., 1984), carrying organic and inorganic pollutants that may pose a severe pollution threat

* Corresponding author. Key Laboratory for Solid Waste Management and Environment Safety, Ministry of Education of China, Tsinghua University, Beijing 100084, China. Tel./fax: þ86 10 6277 3693. E-mail address: [email protected] (D. Yue). 0301-4797/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jenvman.2013.05.014

to both microbial activities and surrounding environment (Mor et al., 2006). For example, a certain proportion of toxic heavy metals in deposited wastes, which can disrupt microorganisms (Bååth, 1989; Giller et al., 1998) would be largely mobilized with dissolved organic matter (e.g., volatile fatty acid) in leachate during initial acidification period (pH 5e6) (Qu et al., 2008; Chai et al., 2007). Serving as hydraulic barriers (Albright et al., 2006), soil covers with low permeability can decrease the filtration of leachate through waste and reduce the migration rate of toxic pollutants, including heavy metals (Yanful et al., 1988a; Navia et al., 2005). Placement of an intermediate cover of soil over every waste cell is required in the national technical guideline in China (Ministry of Construction of the People’s Republic of China, 2004) as well as in many other countries, and the thickness of the cover is recommended to be more than 30 cm. Chinese MSW has been characterized with high organic and moisture content (Wang and Nie, 2001; Zhang et al., 2010) since kitchen waste makes up the highest proportion at approximate 60% (Yuan et al., 2006). Particularly, moisture is always varied from 30 to 60% vs. 20e30% in the U.S. and European countries (Cheng and Hu, 2010; Meng et al., 2012) sometimes even up to 70% (on wet basis) (Wang and Nie, 2001). Hence, leachate, a

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certain proportion of which is from landfilled MSW, is produced in huge volume in Chinese landfills every year. In this case, intermediated soil covers as hydraulic barriers inside the landfill would make significance in transport and distribution of both leachate and accompanied organic and inorganic pollutants, and the biodegradation and stabilization process of solid wastes would be subsequently affected. However, the comprehensive impact of intermediate soil covers on waste stabilization in landfills with extremely high moisture content was rarely investigated. While simulated experiments in laboratories are vulnerable to feed materials and experimental conditions and cannot reflect actual condition, field excavation of landfilled solid waste is capable of revealing the real situation inside a landfill. Therefore, waste samples of different landfill ages were excavated from an old landfill site in China and were characterized and profiled in terms of physical composition, moisture, organic matter, humic substances (HSs), toxic inorganic elements, and microbial community dynamics, to examine the function of intermediate soil cover on landfilled MSW degradation. Based on the above investigation data, both the impact of intermediate soil cover on MSW biostabilization and possible key factors controlling the decomposition degree of MSW were identified and discussed. 2. Materials and methods 2.1. Solid waste sampling and preparation Solid waste sampling was conducted at a sanitary landfill site located in Beijing, China in May 2010. The landfill site, operated from 1996 to 2008, annually received 350,000 metric tons of wastes consisting of MSW and a small proportion of commercial and industrial waste. The final height of the closed landfill was 30 m. It was a typical anaerobic sanitary landfill designed with artificial lining, leachate drainage, and biogas collection systems. In particular, only one 50 cm-thick intermediate soil cover (sandy clay) existed at depth of approximately 19 m from the top of landfill. Above the intermediate cover (from the top to the depth of 19 m), wastes were received from 2004 to 2008. Meanwhile, below the intermediate layer (from the depth of 20 me25 m), landfill ages of wastes were more than seven years. To eliminate heterogeneity of MSW, solid waste samples were collected from four boreholes, 25 m deep each, designated as I, II, III, and IV. In addition, the correlation between landfill ages and landfill depths was established based on landfill operation records. The collected solid wastes were preserved in self-sealing plastic bags immediately before being transferred to laboratory for analysis. Waste samples of 6e10 kg each were collected at depths of 5e 6 m, 8e9 m, 11e12 m, 17e18 m, 21e22 m, and 24e25 m corresponding to 3-, 4-, 5-, 6-, 8- and 10-year-old wastes, respectively. Duplicate samples were collected for each depth. Small proportions of solid wastes (approximate 200 g) from borehole IV, used as the basis in analyzing the microbial community structure, were preserved in self-sealing plastic bags and stored at 4  C before being transported to the laboratory, and the samples were analyzed within 24 h. In order to assure the representative of subsamples for further analysis, the four excavated solid wastes from the same depth of different holes were well mixed and sampled in laboratory according to “Sampling and Analysis Method for Domestic waste (CJ/T 313-2009)” (MOHURD, 2009) to obtain the first-degree sample (<200 mm). 2.2. Analytical methods The excavated solid wastes were characterized in terms of physical composition, moisture content, organic matter [total

organic carbon (TOC), total organic nitrogen (TON), and leachable organic carbon (leachable OC)], HSs [including humic acid (HA) and fulvic acid (FA)], and toxic inorganic elements (including heavy metals Zn, Mn, Cu, Cd, Cr, Ni, Pb and toxic metalloid As). 2.2.1. Physical composition of solid waste The wastes were dried at 50  C in an electric-blast drying oven for 48 he72 h prior to analysis. The solid wastes were then sorted into ash and fines (<20 mm), animals, plants, tiles and ceramics, papers, textiles, glasses, metals, bamboos, and others according to “Sampling and Analysis Methods for Municipal Domestic Refuse” (MOHURD, 2009). The percentage of each fraction was calculated on a dry weight basis. 2.2.2. Moisture content Approximately 150 g of solid waste was randomly collected from sub-samples and was dried under 105  5  C for 24 h. The moisture content was determined from the difference between the initial and dried waste. Each sample was collected in duplicate. 2.2.3. Organic matter content The well-mixed and homogenized waste was first dried in an electric-blast drying oven at 50  C for approximately 48 h. The bulk materials were removed, and the remaining dry waste was then lightly ground to less than 0.3 mm size. The analysis issues were as follows: TOC was determined by oxidizing the solid wastes with potassium dichromate (Chan et al., 2001). TON was obtained by subtracting the ammonia nitrogen (NH3e N) and nitrate nitrogen (NO3eN) from the total nitrogen (TN); TN was determined using the Kjeldahl method. NH3eN was determined using sodium chloride extraction in combination with the sodium hypochlorite colorimetric method, and the NO3eN content was obtained by sodium chloride extraction combined with phenol disulfonic acid colorimetric method (Markus et al., 1985). The VS content of the solid wastes was obtained from the weight loss of approximately 10 g of samples burned in the muffle furnace at 550  C for 2 h (on plastic-free basis). The leachable OC was determined using leaching test. The waste was rotary-oscillated with distilled water at a ratio of 1:10 (w/v) for 24 h, and the suspension was then centrifuged. The leaching procedure was repeated twice, and the combined supernatant was passed through a 0.45 mm filter to remove the particles. Finally, the filtrate DOC was measured using a TOC analyzer (TOC-5000A, Shimadzu, Japan). 2.2.4. HSs content The content of HSs in excavated solid wastes was determined according to the procedure described by Fukushima et al. (2009) with minor modification. Dry waste (dried at 50  C, bulk materials were removed, and lightly ground to pass through 20-mesh sieve prior to use) was extracted using aqueous alkaline solution (0.1 M NaOH þ 0.1 M Na4P2O7, v/v ¼ 1/1) under a N2 atmosphere (m/v ¼ 1/20). After being shaken for 24 h, the suspension was centrifuged and the supernatant was filtered. A 5 mL aliquot of the filtrate was adjusted to pH 1 using concentrated HCl, stirred overnight, and filtered through a 0.45 mm filter using a 2.5 mL aseptic syringe. The precipitate (HA) was washed with distilled water, dissolved using 0.1 M NaOH, neutralized, and was measured DOC (TOC-5000A, Shimadzu, Japan). The supernatant, combined with the distilled water for precipitate washing, was passed through a mini-column packed with 1 mL DAX-8 resin to adsorb FA. Afterward, the column was washed with distilled water until effluent was colorless, and FA was then desorbed, collected and neutralized to measure DOC. The HA and FA contents were calculated and expressed as gram C per kilogram waste on a dw basis, whereas the

G. Qi et al. / Journal of Environmental Management 128 (2013) 259e265

HSs percentage was calculated as percentage of HA and FA in terms of TOC in the solid wastes.

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3. Results 3.1. Physical composition of solid waste

2.2.5. Content of toxic inorganic elements Approximately 0.2 g of each calcined (550  C, 2 h) and ground solid waste (<0.35 mm) was microwave-digested under 195  C for 20 min using combined aqua regia and concentrated Hydrofluoric Acid (HF). The digested solution was passed through a 0.22 mm glassfiber filter and diluted before analysis. The content of toxic inorganic elements was determined using ICP-AES (Zn, Mn, and Cu) and ICP-MS (As, Cd, Cr, Ni, and Pb). The ICP-AES analyses were carried out on an IRIS Intrepid II XSP (ThermoFisher, USA) instrument, and the ICP-MS analyses were carried out on a Thermo ICP-MS XII (ThermoFisher, USA) instrument. All samples were processed in duplicate. 2.3. Microbial community dynamics The microbial community dynamics of the solid wastes from borehole IV were investigated using specific polymerase chain reaction (PCR) followed by terminal-restriction fragment (T-RF) length polymorphism (T-RFLP) of 16S rRNA gene, and the main procedure was described by Wang et al. (2010). For T-RFLP analysis of methanogenic archaea, the primer pairs used for amplifying were 21F (50 -TCCGGTTGATCCTGCCGGA-30 ) and 1495R (50 CTACGGCTACCTTGTTACG-30 ). Moreover, the purified PCR products were digested using a certain volume of 10  Taq I Buffer (TaKaRa) and 10  BSA (TaKaRa) at 65  C for 6 h. 2.4. Data treatment 2.4.1. Solid-waste characterization All quantitative data on waste physicochemical indexes and heavy metal contents were calculated from the average value of the samples from the four sampling boreholes (I, II, III, and IV) at the same depth, and the standard derivation (SD) was also calculated to determine waste heterogeneities. 2.4.2. Microbial community dynamics With regard to the microbial community dynamics characterized by specific PCR in combination with T-RFLP, the relative T-RF abundances were calculated according to the method of Wang et al. (2010), expressed as a proportion of a given peak area to the total selected peak area of each sample. Peaks with an area of less than 1% of the total were excluded, and the percentage of each remaining peak was recalculated until all peaks have relative abundance of more than 1%. The T-RF number, ShannoneWeiner’s index (H0 ), and Paretoe Lorenz (PL) evenness curves were introduced to evaluate the microbial richness, microbial diversity, and functional organization of microbial community (interspecies distribution) inside the landfill, respectively (Marzorati et al., 2008; Wittebolle et al., 2008). H0 was calculated from

H0 ¼


n X

Visual inspection of the excavated waste material with dark color and strong ammonia smell indicated anaerobic conditions inside the landfill site. The physical composition of the excavated wastes showed that the major identifiable components of the waste were ash and fines (<20 mm), plastics, and ceramics in the range of 73e89% (w/w), 3e10% (w/w), and 0.4e5% (w/w), respectively (data shown in the Supplementary material). Compared with the original waste characterized with high organic matter content (30.1e66.2%, on a wet basis) and high moisture content (45.2e65.5%), visible distinguishing organic fractions such as plant and animal in the incipient MSW, which were mainly from kitchen waste, were not identifiable. However, no statistical significant difference in the physical composition of solid-waste variation with landfill age was observed. 3.2. Spatial distribution of moisture and toxic inorganic elements The spatial distribution of moisture content at depth of 5e25 m is shown in Fig. 1. The values for all waste samples varied between 25.9% and 40.7%, much lower than the optimum value (50e60%) for methanogenesis (Kelly et al., 2006). Interestingly, the highest moisture content appeared at the depth of 14e18 m, which was 2.3e2.4 times that in the underlying intermediate soil layer (16.7% moisture). Meanwhile, the lowest moisture content appeared in the 8- and 10-year-old wastes (25.9% and 30.2%, respectively), which were lying below the intermediate soil cover (at 19 m in depth from the top of landfill). In this case, the moisture may have infiltrated through the unsaturated and porous solid wastes and transported to the underlying strata above the compacted soil cover, where moisture accumulated because of the relatively low permeability of the intermediate soil layer, as reported previously (Korfiatis et al., 1984; Whitworth and Ghazifard, 2009). Without moisture supply from the upper strata and experiencing continuous drainage by the bottom leachate collection systems, the MSW below the intermediate soil layer exhibited the lowest moisture content. Accordingly, the moisture distribution exhibited two different patterns because of the compacted intermediate soil cover barrier. Table 1 shows the concentrations of the eight most common toxic inorganic elements in solid wastes. Zn, Mn, and Cu, whose

Pi lnPi

i¼1

where Pi is the relative abundance of T-RF (i), and n is the total number of selected T-RFs. To interpret numerically the PL curves, the 45 diagonal corresponding to the perfect evenness of a community was selected to evaluate the microbial structure dynamics. The more the PL curve deviates from the 45 theoretical perfect evenness curve, the lesser is the observed evenness in the structure of the studied community, and a smaller fraction of different species is present in dominant numbers.

Fig. 1. Spatial distribution of the moisture content of excavated solid wastes. Dashed line indicates the location of the intermediate soil cover, whose moisture content is 16.7%.

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Table 1 Concentrations of toxic inorganic elements in excavated solid wastes (mg kg
1 on dw basis). Landfill age. (yrs)

Depth (M)

As

3 4 5 6 8 10

5e6 8e9 11e12 17e18 21e22 24e25

7.44 5.69 8.84 10.1 6.24 7.85

Cd (3.02)a (2.31) (2.20) (3.7) (2.23) (3.88)

0.31 0.28 0.43 0.98 0.26 1.36

Cr (0.16) (0.14) (0.32) (0.25) (0.22) (1.89)

54.89 57.10 45.47 85.49 62.22 54.13

Cu (4.08) (34.18) (5.70) (50.07) (18.86) (9.40)

156.0 100.1 116.5 259.7 80.6 89.2

Mn (76.3) (32.3) (37.4) (83.1) (22.7) (46.1)

551.7 521.5 695.7 530.4 510.0 494.5

Ni (47.2) (52.5) (282.3) (32.1) (25.4) (72.8)

19.8 18.3 24.5 30.5 20.0 18.8

Pb (2.2) (1.8) (12.8) (9.9) (3.0) (2.7)

51.4 57.5 50.7 84.0 48.0 55.8

Zn (15.6) (18.1) (7.0) (39.4) (12.7) (14.6)

349.5 440.6 650.0 981.7 371.7 341.9

(80.1) (167.6) (442.1) (516.9) (112.7) (152.5)

Note: a Data in parentheses are SD.

average content varied from 341.9 to 981.7 mg kg
1 dw, 494.5e 695.7 mg kg
1 dw, and 80.6e250.7 mg kg
1 dw, respectively, were the main toxic heavy metal species. The high contents of Cu and Zn probably came from kitchen waste, ash, plastics, and paper components of the incipient waste (Long et al., 2011). All analyzed contents of toxic inorganic elements were of different degrees lower than those in a landfill in East China (Chai et al., 2007; He et al., 2006) with unknown reason. Similar to moisture, the highest content of toxic inorganic elements also appeared in 6-year-old waste stratum (except for Mn). It might be conjectured that toxic inorganic elements were considerably mobilized from the upper waste in the early acidification period and migrated downward with dissolved organic compounds (e.g., short-chain volatile fatty acid) (Abu-Rukah and AbuAjarayesh, 2002; Li et al., 2009; Loch et al., 1981; Qu et al., 2008; Yanful et al., 1988b). However, their mobility was attenuated upon meeting the compacted intermediate soil cover (sandy clay) with low permeability (Mimides and Perraki, 2000). A small proportion of toxic inorganic elements (mainly heavy metals) precipitated mostly as secondary carbonates or hydroxides at the interface (Yanful et al., 1988b) or adsorbed on the intermediate cover (Loch et al., 1981), whereas most toxic inorganic elements accumulated in the waste just above the intermediate cover. Below the intermediate soil cover, however, concentrations of toxic inorganic elements in 8- and 10-year-old waste were nearly equal to that of 3and 4-year-old waste, and it would be due to migration of certain proportion of toxic inorganic elements dissolved in leachate through drainage systems in the early acidification period. 3.3. Variation in the remaining organic matter The TOC, TON, and VS contents were used to quantify the remaining organic fraction of MSW (Fig. 2). The TOC, TON, and VS

mean values in the excavated MSW varied from 50.1 g kg
1 dw to 83.3 g kg
1 dw, 2.0 g kg
1 dw to 5.1 g kg
1 dw, and 5.3%e10.2% (w/ w), respectively. The relatively low organic matter content in the excavated wastes in comparison with incipient MSW indicated biodegradation of the deposited waste and the co-disposal of construction waste in this landfill. A dramatically high amount of organic matter remained in the 6-year-old wastes, just above the intermediate soil cover (TOC ¼ 83.3 g kg
1 dw, TON ¼ 5.1 g kg
1 dw, and VS ¼ 10.2%). Moreover, the TOC, TON, and VS values in the 8- and 10-year-old wastes were larger than expected, approximately equal to that of the 3- to 5-year-old waste, which ranged from 50.5 g kg
1 dw to 52.8 g kg
1 dw, 2.0 g kg
1 dw to 3.5 g kg
1 dw, and 5.3%e7.0%, respectively. In contrast to theoretical knowledge (Francois et al., 2006; He et al., 2011), the remaining organic matter did not decrease with increasing landfill age. The leachable OC (expressed as percentage of TOC in solid wastes) of the waste was used to determine whether the remaining organic matter is from the upper waste strata (e.g., by moisture transport) or just an innate residue. If the value of the leachable OC is relatively high, the organic matter comes mainly from upper waste strata, whereas a low leachable OC value would indicate that the organic matter is mainly an indigenous residue. As shown in Fig. 2, the leachable OC of the excavated waste gradually decreased from 3.6e4.6%e1.2% with the increase in landfill age, dramatically lower than that of the intermediate soil layer (10.8%), indicating that the TOC fraction of the wastes were mainly innate residue. The abnormally high TOC fraction content in the 6-year-old waste and comparatively high content of remained organic matter in 8- and 10-year-old waste might be resulting from composition fluctuation of incipient MSW or because the decomposition and/or transformation of native organic matter had been inhibited for some reasons (e.g., low microbial activity). 3.4. Evolution of HSs content

Fig. 2. Organic matter content variation of the excavated solid wastes. The TOC, TON, VS, and leachable OC in the intermediate soil cover are 10.0 g kg
1 dw, 1.2 g kg
1 dw, 3%, and 11%, respectively.

Humification is considered as a stabilization stage after the intense biodegradation and mineralization of organic matter, and HSs could have significantly been formed during this period. To evaluate the stabilization degree of the remaining organic matter in the solid wastes (Alburquerque et al., 2009; Dias et al., 2010; Sellami et al., 2008; Zhu and Zhao, 2011), the absolute contents of HA and FA (expressed as amount of TOC in solid waste on dw basis) were measured, and the HSs relative percentage was calculated (Fig. 3). The absolute HA and FA contents varied from 5.5 to 13.7 g C kg
1 and from 5.8 to 9.3 g C kg
1, respectively. The highest HA and FA contents appeared in 6-year-old waste, and this corresponded to the highest organic matter content in this stratum. The percentage of HSs coarsely decreased from 42% in the 3-year-old waste to 26% in the 10-year-old waste except for the 8-year-old waste, indicating that humification of solid wastes proceeded not well despite being exposed to longer landfill time (especially 6-year-old waste and 10-

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263

Fig. 3. HSs content in the excavated solid wastes. The absolute HA content, absolute FA content, and percentage of HSs in the intermediate soil cover are 2.1 g C kg
1 dw, 1.2 g C kg
1 dw, and 3%, respectively.

year-old waste). The relatively high HSs percentage in the 3- and 8year-old wastes can be attributed to the humification process enhanced by large amount of metal cation leaching from the upper soil cover (Bosetto et al., 2002). In particular, for the 3-year-old waste, little oxygen may have been provided an access to migrate from the atmosphere to the top waste stratum, which is beneficial for the production of high molecular HSs (Sinha, 1972), and consequently, the humification degree was the highest. 3.5. Microbial community dynamics As seen from the PL evenness distribution curve (Fig. 4), 20% of the T-RFs corresponded to 66e72% of the cumulative T-RF relative abundance for bacteria, 20% of the T-RFs corresponded to 65%e83% of the cumulative T-RF relative abundance for methanogenic archaea. The internal structure of the bacteria and methanogenic archaea communities indicated that only a small group of species played a numerically dominant role in both waste decomposition and methane production, and both bacterial and methanogenic archaea communities in the landfill were highly functionally organized (Briones and Raskin, 2003). Such highly organized microbial ecosystems in function indicated the highly selective environment inside the landfill. Table 2 shows the number of T-RFs and the ShannoneWeiner index (H0 ) for both bacteria and methanogenic archaea. The bacteria and methanogenic archaea exhibited the lowest values of T-RF numbers and H0 in the 6-year-old waste stratum corresponding to 13 and 1.8 for bacteria and 8 and 1.4 for methanogenic archaea, indicating the lowest microbial activity (survival rate) and microbial diversity in the 6-year-old waste. The microbial ecosystem in this stratum experienced large interference and damages. It was noticeable that the number of T-RFs and the ShannoneWeiner index (H0 ) for both bacteria and methanogenic archaea in 10-year-old waste were the highest, but organic matter in this stratum remained unstable. This would attribute to function and metabolism inhibition of certain microbial species for some reason (e.g., lack of moisture). 4. Discussion 4.1. Barrier effect of the intermediate soil cover The intermediate soil cover in the investigated landfill was 50 cm thick, and its negative effect as a real barrier was obvious.

Fig. 4. PL evenness distribution curve of the methanogenic archaea and bacteria TRFLP profiles. The dashed vertical line at the 0.2 x-axis level is plotted to evaluate the range of the Pareto values.

Both moisture and toxic inorganic elements accumulated dramatically in the 6-year-old waste, which was directly over the intermediate soil layer, whereas the moisture and heavy metal content in solid wastes below the soil layer were much lower. The lowpermeability intermediate soil cover not only created barriers that stop the leachate and water from percolating (Miller et al., 1991) but also provided a possibility for the retention of toxic inorganic elements as previously reported (Mimides and Perraki, 2000).

Table 2 T-RF numbers and H0 for the bacteria and methanogenic archaea in solid waste with respect to landfill depths and landfill ages. Landfill age (yrs)

Depth (m)

3 4 5 6 8 10

5e6 8e9 11e12 17e18 21e22 24e25

Bacteriab

Methanogenic archaeac

T-RFs

ShannoneWeiner index

T-RFs

ShannoneWeiner index

ea 34 22 13 29 51

e 2.7 2.2 1.8 2.4 3.2

18 7 e 8 22 22

2.2 1.2 e 1.4 1.5 2.2

Notes: a Means not analyzed. b The T-RF numbers and ShannoneWeiner index for bacteria in intermediate soil cover are 20 and 1.9. c The T-RF numbers and ShannoneWeiner index for methanogenic archaea are 8 and 1.1.

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The investigated landfill was divided into two different sublandfills because of the barrier effect of the intermediate cover. The upper stratum of solid wastes functioned more like a “controlled tipping” system with low-permeability ground layer but no leachate collection system, and the lower stratum acted as a contained landfill system (Cossu, 2010). Hence, two types of landfilling appeared in one single landfill. Many currently running landfills in China are using high-density polyethylene (HDPE) membrane instead as intermediate cover so that the landfill space can be saved. However, lots of HDPE membrane is still kept inside waste piles when disposing of MSW again at the same area. Quite negative barrier effect could be imagined according to the results of this investigation. 4.2. Biodegradation of organic fraction of MSW in landfills The degradation and stabilization of landfilled MSW are largely dependent on the metabolism of various microorganisms (especially bacteria and methanogenic archaea) and directly related to the microbial community structure and functional organization. The investigation results showed that the 6-year-old waste had the lowest microbial species (T-RF numbers) and the lowest microbial activity and diversity (H0 ) despite the highest moisture content in the solid waste. Therefore, more organic matter (35.4e66.3%) remained not degraded or transformed. It is noteworthy that a comparatively high amount of toxic inorganic elements existed in the 6-year-old waste in comparison with the other waste samples, especially Zn and Cu. The Zn content was 3.3 times the current mandatory European Union (EU) limit and Chinese legislation (300 mg kg
1 soil maximum), whereas Cu was 1.9 times the EU mandatory limit (140 mg kg
1 soil maximum) (Ghorbani et al., 2002) and was 2.7 times the quality standard limit to protect human health in China (100 mg kg
1 soil maximum). As for the other detected toxic inorganic elements (Ni, Cd, Cr, Pb, and As), the concentrations were all lower than the corresponding EU limit or environmental quality standard for soils in China (Table 1 in supplementary material). High concentrations of Zn and Cu have been determined to decrease the turnover rate of organic matter in soil (Chander and Brookes, 1991): The Cu content of the soil at Luddington, which contained 32% more organic matter than an uncontaminated soil, was 3.7 times the current EU mandatory limits; The soil at Lee Valley, which contained 10% more organic matter than the control samples, was contaminated with Zn at 3.4 times the permitted concentration. In addition, this investigation identified highly positive correlations between Zn and the remaining organic matter (R2 ¼ 0.64) and that between Cu and the remaining organic matter (R2 ¼ 0.81) (data not shown). In addition, above two heavy metals were coarsely negatively correlated to the microbial communities. Therefore, the high concentrations of toxic inorganic elements (especially Zn and Cu) in the 6-year-old waste were primarily responsible for the observed reductions in microbial activities and the highly functional organization in the microbial communities (Kuperman and Carreiro, 1997), finally leading to the significantly reduced turnover rate of organic matter. As previously discussed, the biodegradation of landfilled MSW above the intermediate cover can be assumed as follows: the downward water transport in the early landfill acidification period mobilizes the toxic inorganic elements (especially Zn and Cu), as well as the volatile fatty acids and other organic compounds in the waste. However, due to the barrier effect of the intermediate soil cover, toxic inorganic elements accumulate in the underlying strata above the cover, causing subsequent biological toxicity to the microbes and finally leading to insufficient decomposition and/or stabilization of the underlying landfilled MSW (6-year-old waste), even for several decades.

Below the intermediate layer, further inefficient decomposition and stabilization of the solid wastes (especially 10-year-old waste) were also observed despite the comparatively low toxic inorganic element contents, longer landfill time (eight years), and more microbial species and diversity. In this case, the actually low moisture content may become one of the key factors inhibiting microbial function and metabolism and subsequently slowing down the biological degradation and stabilization of solid wastes. The biodegradation and stabilization processes of the MSW organic fractions inside the landfill are actually complex because various environmental stress factors, such as toxic inorganic elements, moisture, and landfill operating mode and design (e.g., material type of daily and intermediate cover), affect the functional organization of microbial ecosystem and organic matter decomposition subsequently. The decomposition and stabilization process of solid wastes characterized in this investigation is an excellent case to illustrate the aforementioned complexity. 5. Conclusion The characteristics of wastes with different landfill ages in an old landfill were profiled to discuss the impact of intermediate soil cover on landfill stabilization. Moisture and toxic inorganic elements largely accumulated, and a relatively high amount of unstable organic matter remained in the 6-year-old waste underneath which lay a compacted intermediate soil cover. The Cu or Zn concentration was highly positively correlated to the residual organic matter content but coarsely negatively correlated to microbial communities. The organic matter in 8- and 10-year-old wastes lying under the intermediate cover were not as well degraded as expected, possibly because of the lack of water resulting from barrier effect of the compacted intermediate soil cover. The accumulation of toxic inorganic elements over the intermediate cover led to the reduction in microbial activities, which in turn resulted in relatively low biodegradation of organic matter. On the contrary, the lack of water caused slow waste degradation below the intermediate soil layer because of weakened microbial activities. Acknowledgment This project is supported by special fund of State Key Joint Laboratory of Environment Simulation and Pollution Control (No. 11Z02ESPCT) and the National Natural Science Foundation of China (No. 51208280). Appendix ASupplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.jenvman.2013.05.014. References Abu-Rukah, Y., Abu-Ajarayesh, I., 2002. Thermodynamic assessment in heavy metal migration at EI-Akader landfill site, North Jordan. Waste. Manag. 22, 727e738. Albright, W.H., Benson, C.H., Gee, G.W., Abichou, T., Tyler, S.W., Rock, S.A., 2006. Field performance of three compacted clay landfill covers. Soil. Sci. Soc. Am. J. 5, 1157e1171. Alburquerque, J.A., Gonzálvez, J., Tortosa, G., Baddi, G.A., Cegarra, J., 2009. Evaluation of “ alperujo” composting based on organic matter degradation, humification and compost quality. Biodegradation 20, 257e270. Bäumler, R., Kögel-Knabaner, I., 2008. Spectroscopic and wet chemical characterization of solid waste organic matter of different age in landfill sites, Southern Germany. J. Environ. Qual. 37, 146e153. Bååth, E., 1989. Effects of heavy metals in soil on microbial processes and populations (a review). Water. Air. Soil. Pollut. 47, 335e379. Bilgili, M.S., Demir, A., Özkaya, B., 2007. Influence of leachate recirculation on aerobic and anaerobic decomposition of solid wastes. J. Hazard. Mater. 143, 177e 183.

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