The use of the dynamic respiration index to predict the potential MSW-leachate impacts after short term mechanical biological treatment

Bioresource Technology 128 (2013) 351–358

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The use of the dynamic respiration index to predict the potential MSW-leachate impacts after short term mechanical biological treatment Silvia Salati a, Barbara Scaglia a, Alessandra Di Gregorio b, Alberto Carrera b, Fabrizio Adani a,⇑ a b

RICICLA GROUP, Dipartimento di Scienze Agrarie e Ambientali: Produzione, Territorio, Agroenergia, Via Celoria 2, 20133 Milan, Italy SORAIN CECCHINI TECNO Srl, Via di Pontina 545, 00128 Roma s.r.l., Italy

h i g h l i g h t s " We aimed to describe leachate impact reduction after short-term full scale MBT process. " The dynamic respiration index (DRI) was used to measure leachate impact reduction. " Leachate impact was measured after 1 year in landfill reactors. " Short-term MBT getting a DRI < 1000 mg O2 kg VS " COD, BOD5 and total heavy metal reduced of

a r t i c l e

i n f o

Article history: Received 22 August 2012 Received in revised form 19 October 2012 Accepted 23 October 2012 Available online 31 October 2012 Keywords: Dynamic respiration index Landfill impacts Leachate Mechanical–biological treatment Municipal solid waste

1

70%,

h

1 greatly reduced leachate impacts. 81% and 16%, respectively.

a b s t r a c t The aim of this work was to evaluate the effects of short full scale MBT process (28 d) getting a biological stability of DRI < 1000 mg O2 kg VS 1 h 1, on the impact of leachate produced in simulated landfill. As consequence of that, waste was processed by full scale MBT and both untreated and treated waste were successively incubated in simulated landfills for 1 year. Leachates were collected at different incubation times and characterized. Results obtained indicated that MBT allowed waste-organic matter (OM) reduction favoring, also, optimal condition for successive OM degradation in the simulated landfill. Final results indicated a total reduction of leachate impact for the treated waste (DRI of 978 mg O2 kg VS 1 h 1) with respect to the untreated waste of: 54%, 69%, 77%, 70%, 81% and 16% for NTK, NH3, TOC, COD, BOD5 and total heavy metal contents, respectively. Ó 2012 Elsevier Ltd. All rights reserved.

1. Introduction Municipal solid wastes (MSW) are mainly disposed into landfill, which represents the simplest and cheapest method of waste management (Allen, 2001; Zheng-Hao et al., 2009). Nevertheless MSW disposal in landfill caused different impacts, i.e. methane and other greenhouse gases and hazard emissions, odours production, polluted leachates, etc. In landfill the biodegradation of organic fractions contained in MSW is the major mechanism that governs biological pollution with particular reference to leachate productions (Read et al., 2001). Leachate is formed when moisture content in waste exceeds its water holding capacity (WHC), i.e. maximum moisture that is retained in a porous medium without producing percolation (Driessen et al., 1995; El-Fadel et al., 1997). Water percolation provides a medium in which waste can undergo into simpler ⇑ Corresponding author. Tel.: +39 0250316544; fax: +39 0250316521. E-mail address: [email protected] (F. Adani). 0960-8524/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2012.10.101

substances through a range of chemical reactions (dissolution, hydrolysis and redox reactions) and microbial metabolism activity (El-Fadel et al., 1997). As result of these processes, leachates contain organic molecules, inorganic water-soluble component, heavy metals and xenobiotic compounds (Kjeldsen et al., 2002). The chemical composition of leachates varies in function of landfill age (Li et al., 2010), MSW biological stability and moisture content (Barlaz, 2006). Therefore leachate characterization is often used as an indicator of the landfill conditions and it is essential to understand the long-term effects of landfills (Kjeldsen et al., 2002). Young landfill leachates are characterized by high organic carbon concentrations (COD of 10,000 mg O2 l 1), high content of biodegradable fraction (BOD5/COD in the range 0.4–0.7) and low pH values (pH < 6.5), whereas leachate of old landfill shows lower organic carbon content (COD < 4000 mg l 1; BOD5/COD < 0.1) and higher pH value (7.5–8.5) than young landfill. In order to reduce or prevent environmental pollution, the European Commission emanated the Landfill Directive (European Parliament and Council Directive, 1999) to drive the member states

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to draw up strategies for progressively reducing the amount of biodegradable MSW in landfill. The reduction of organic fraction of MSW destined to landfill can be obtained by three different approaches: (i) source separated collection of organic fraction of MSW to produce compost; (ii) MSW burning to produce energy and, (iii) mechanical–biological treatment (MBT) of MSW to produce a stabilized or a composting-like material (Scaglia and Adani, 2008). MBT consists of mechanical pre-treatment of MSW followed by an aerobic (composting-like process) or anaerobic/aerobic degradation process (Scaglia et al., 2010). During the process, the labile organic fractions (i.e. simple lipids, proteins, cellulose and hemicelluloses) are degraded (Scaglia et al., 2010), while, more recalcitrant molecules are preserved and concentrated (Scaglia et al., 2010). The length of the biological treatment can vary, so that a biostabilized product shows different characteristics. The measure of the efficacy of the aerobic process consist in a direct or indirect measurement of the decomposition degree of the easily degradable OM vs the presence of recalcitrant OM. The biological stability measures the extent to which readily biodegradable organic matter has been decomposed (Scaglia and Adani, 2008). The most used and accepted methods to measure the biological stability are the respirometric methods (Scaglia and Adani, 2008). Among them, the Dynamic Respiration Index (DRI) offers the possibility to reproduce full scale conditions in laboratory, because sufficient amount of waste can be used to preserve the self heating condition (Scaglia et al., 2011). In addition, DRI was officially recognized at EU level becoming standardized (European Committee for Standardization, 2011) and validated methods (Scaglia et al., 2011). Nevertheless, new research has been, recently, performed in order to improve this methodology (e.g. Komilis and Kanellos, 2012) or to proposed simplified approach (Cossu et al., 2012). By now, no many data about the impact of biologically treated waste in landfill are available with particular references to shortterm MBT treatment (2–4 weeks) and with the use of DRI as an index to predict potential waste impacts. In addition only indirect measurements of landfill impacts of biostabilized waste characterized by DRI below the limits proposed by Italian legislation and international literature (Scaglia et al., 2010) (DRI < 1000 mg O2 kg VS 1 h 1), are available. The aim of this work was to evaluate the effects of full scale short-term MBT process (28 d) getting a biological stability of DRI < 1000 mg O2 kg VS 1 h 1, on the impact of leachate produced in simulated landfill. Doing so untreated and treated wastes coming from the full scale MBT plant, were disposed in simulated landfills for 1 year, leachates collected and fully characterized. 2. Methods

of wet weight (w.w.) of untreated and biostabilized MSWs were brought to the laboratory. Samples were stored at 4 °C and processed within 3 to 5 days from receipt. A homogeneous sub-sample of 3 kg was taken from each USMSW sample to determine dry matter (DM) content after sample drying at 105 °C. Successively, samples were reduced in size (1 mm of diameters) and used to determine volatile solid (VS), total organic carbon (TOC), total Kjieldal nitrogen (TKN) and ammonia–nitrogen (N–NH3) (on wet samples) contents, and water holding capacity (WHC), according to The US Composting Council methods (The US Composting Council, 1997). In order to detect heavy metal (HM) contents USMSW samples were previously digested with concentrated nitric acid (EPA, 2007). Then HMs were determined by using an inductively coupled plasma mass spectrometer (EPA, 2007) (ICP–MS, Varian, Fort Collins, USA). A certified standardized material (GBW 07405, soil – National Centre for Standard Materials, Beijing, China) was used as analytical reference. All analyses were performed in triplicate. Biological stability of USMSW samples was determined by detecting the Dynamic Respiration Index (DRI) (European Committee for Standardization, 2011; Scaglia et al., 2010) using a 30 L adiabatic respirometric reactor (Costech International, Cernusco S.N., Italy). 2.3. Landfill lab-scale reactors and MSW incubation No-biostabilized and biostabilized USMSW samples were incubated at laboratory-scale landfill reactors for 12 months. Landfill reactors consisted in prototype Plexiglass reactors (high of 150 cm and Ø of 25 cm) (Fig. S1). Reactors were designed to allow rainfall simulation from the top of the reactor and to collect leachates from the bottom. Lab-scale reactors were loaded with 8.8 kg w.w. of I-USMSW and S-USMSW at a WHC of 75% (Scaglia et al., 2010), resulting a final bulk density of the mass of 0.8 Mg m 3. Reactors were hermetically sealed and flushed with N2 for 2 h before their closure. Anaerobic conditions were periodically verified using anaerobic kit test (microbiology anaerotest, Merck, NJ 08889-0100 USA). The trials started on April 2010 and finished on March 2011 (12 months length). Three replicates were performed for each USMSW (three reactors), for a total of six reactors. 2.4. Leachate production and characterization 2.4.1. Rain events Artificial rain (amount and event periods) was reproduced taking into consideration rain events measured in Rome Ciampino location (1970–2007 series) such as reported by the European Climate Assessment & Dataset (http://eca.knmi.nl/dailydata/) (Table S1), being this location very close to the Rome landfill (Italy) in which biostabilized material under study will be allocated in the next future.

2.1. Biostabilization process Biostabilization process was performed at the MBT full-scale plant of Sorain Cecchini Tecno, located in Rome, Italy, treating 1200 Mg d 1 of unsorted MSW. Biological process was performed to treat the undersize fraction of MSW (USMSW) coming from MSW sieving (sieve-hole diameter of 90 mm); biological process was conducted for 28 d under forced aeration and mass turning (more details on http://www.soraincecchini.it/, on July 12, 2012). 2.2. MSW samples The USMSW was sampled at the start-point of the biological process (I-USMSW, i.e., undersize fraction of MSW sieved at 90 mm) and after 28 d of the biological process (S-USMSW). Samples were took by using standard sampling procedures (European Committee for Standardization, 2006); doing so about 40–50 kg

2.4.2. Leachates collection and characterization Leachates produced by simulated landfills were daily collected and quantified; afterwards, they were mixed together to obtain two-months period samples, i.e., T1 = April–May, T2 = June–July, T3 = August–September, T4 = October–November, T5 = December– January and T6 = February–March, and stored at 20 °C. The following analytical parameters were systematically monitored using APHA methods (1998): pH, volatile fatty acids (VFA), total alkalinity (TA), TOC, COD and BOD5, Cl , F , SO42 and CN (European Parliament and Council Directive, 1999). In addition dissolved organic carbon (DOC) was determined (APHA, 1998) on the solutions obtained after filtration of leachates throughout 0.45 lm Millipore filters (Advantec MFS, Pleasanton, CA). TKN and N–NH3 contents in leachates were performed by using methods reported for wastewater (IRSA CNR, 1994). Metal concentrations in leachates were determined after acid digestion such as reported in Section 2.1.

S. Salati et al. / Bioresource Technology 128 (2013) 351–358

2.5. Statistical analysis All statistical analyses were performed using analysis of variance (ANOVA) with the Tukey test used to compare means (SPSS statistical software, SPSS Chicago IL).

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fact while the treated sample (S-USMSW) showed a decrease of WHC during incubation because of OM degradation, the untreated sample (I-USMSW) showed a strong increase in this parameter probably due to OM modification, i.e. hydrolysis of polymeric molecules and successive oxidation of OM that increased the ability to retain water.

3. Results and discussion 3.1. MSW biostabilization process MBT process determined a degradation of organic matter contained in the waste: TOC and VS contents reduced during the process by about 30 g kg 1 on a relative basis (Table 1). Absolute reduction was of 162 and 119 g kg 1 for TOC and VS, respectively. These results indicated a low degradation rate after MBT process. Nevertheless the high content in the waste of non-biogenic VS (plastic and rubber, i.e. 310 g kg VS 1) masked the VS reduction suggesting that low degradation rate was only apparent and thereby these data were not suitable to adequately describe the effect of MBT on USMSW. Therefore biological stability on waste samples before and after MBT were measured. MBT process led to a stabilization of the waste, such as indicated by DRI (Table 1) that after 28 d of biological process (S-USMSW) was of 978 ± 153 mg O2 kg VS 1 h 1, which fully respected Italian rules for landfill disposal of biological pre-treated waste (DRI < 1000 mg O2 kg VS 1 h 1) (Gazzetta Ufficiale, 2010). Chemical parameters characterizing untreated (I-USMSW) and biologically-treated waste (S-USMSW) are resumed in Table 2. 3.2. Leachate: quali-quantitative characterisation 3.2.1. Quantification of the leachate productions during simulated landfill Leachate productions and their trends during the incubation period resulted similar for both wastes incubated (Table 2). During the first month of incubation (samples T1–T5) the leachate volumes were similar to that of water volume added by simulated rains. Successively, i.e. T6 samples, the leachate volume of IUSMSW was lower than that of simulated rains (Tables S1 and 2), whereas the leachate volume of S-USMSW was much higher than that of the corresponding rain-events. This difference was due to the different ability of waste to retain water (Table 2); in

Table 1 USMSW characterisation. I-USMSW pH DM DRI VS TOC TKN N–NH3 Ni Cd Cu Hg Se Mo Zn Cr Pb Ba As Sb

S-USMSW a

(g kg w.w 1) (mg O2 kg VS (g kg dm 1)

(mg kg dm

1

)

1

h

1

)

5.6 ± 0.2a 44.4 ± 1.56a 3910 ± 358b 556 ± 19a 298 ± 15b 11.7 ± 0.8a 4.28 ± 0.34a 15 ± 0.81a 2.84 ± 0.04a 106 ± 0.5a 3.58 ± 1.12a 0.04 ± 0.00a 1.1 ± 0.23a 127 ± 10a 23 ± 0.6a 219 ± 1.3a 113 ± 2a 2.93 ± 0.27a 0.79 ± 0.03a

6.8 ± 0.1b 47.5 ± 2.03a 978 ± 153a 525 ± 15a 267 ± 8a 11.3 ± 1.9a 5.32 ± 0.74a 16.5 ± 0.07b 3.24 ± 0.10b 116 ± 0.2b 3.94 ± 0.42a 0.07 ± 0.00b 1.37 ± 0.13a 135 ± 1a 25.0 ± 0.2b 237 ± 2b 118 ± 1b 3.23 ± 0.18a 0.92 ± 0.00b

a Means followed in the same line by the same letter are not statistically different (p < 0.05) according to Tukey test.

3.2.2. Chemical characterisation of leachates Simulated landfill affected leachate chemical characteristics (Table 2). The pH values of leachates after two months of incubation in lab-reactors (sample T1) were similar for the two wastes studied (Table 1); nevertheless different trends were observed during the successive incubation months. The pH values of I-USMSW leachates rapidly decreased during first stage of incubation (sample T2) remaining constant until final sampling. Final pH value registered was of 5.9 that is typical of young landfill characterized by an acid phase (Kjeldsen et al., 2002). On the other hand, the pH of the S-USMSW sample increased during first stages of incubation reaching a peak after 6 months of incubation (sample T3), then it decreased to a final value of 8.7, that is typical of an old landfill characterized by a methanogenic phase (Kjeldsen et al., 2002). Volatile fatty acids content agreed with pH data. VFAs content of the I-USMSW increased during first stage of trial, but, starting from the 4th month of incubation (sample T2) it decreased reaching a final content lower than that of starting samples (28.3 and 0.92 g CH3COOH l 1 for I-USMSW and for sample T1 and T6, respectively). The S-USMSW showed a completely different behaviour since VFAs content, that was much lower than that of IUSMSW (Table 2), reduced rapidly after the first 2 months of incubation (Table 2). Total waste-TKN content before incubation was similar for the two wastes studied (sample T1). Incubation in simulated landfill led to a generalized decreasing of TKN contents for both leachates studied with final TKN contents that decreased of about 80% and 95% with respect to the corresponding starting samples (Table 2). Despite to this general behaviour the I-USMSW sample showed an increases of TKN content after 2 months of incubation (Table 2). This increase might be explained considering the strong leaching of soluble-OM containing N (i.e. amino acid) during the first stage of incubation because of hydrolysis of fresh organic matter (Crawford and Smith, 1985). TKN was composed almost by ammonia (Table 2), as the comparison of the two parameters, also, suggested (r = 0.90, p < 0.05, n = 12). Therefore ammonia contents showed the same trend of those indicated for TKN, although NH3 content of I-USMSW, contrarily to that indicated for TKN, did not increase for sample T2. Results showed a strong reduction of the ammonia content in leachates after 1 year of incubation (sample T6) (Table 2) ( 96.6% for I-USMSW and 84% for S-USMSW); nevertheless S-USMSW showed N–NH3 content of about 10-times lower than that of I-USMSW. Final ammonia content were of 0.4 and 0.06 g l 1 for I-USMSW and S-USMSW, respectively. Total alkalinity for both samples incubated decreased during anaerobic incubation because of ammonia depletion. Good correlation found between these two parameters confirmed this observation (r = 0.91, p < 0.05, n = 12). 3.2.3. Biological characterization of leachates To evaluate the biological impact of untreated and biostabilized wastes, TOC, DOC, COD and BOD5 parameters were detected on leachates sampled at different incubation time. Despite starting VS and TOC contents of I-USMSW and S-USMSW were very similar (Table 1) the acquirement of a good degree of biological stability during the aerobic biological process determined a different content of TOC in leachates of wastes studied (Table 2). In particular, after just 2 months of incubation

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Table 2 Chemical and biological characteristics of leachates. Time

I-USMSW

S-USMSW

Production (l)

T1 T2 T3 T4 T5 T6

2.2 ± 0.1a C 0.86 ± 0.2a A 1.6 ± 0.1a B 3.5 ± 0.1a D 2.1a C 1.3a B

2.2 ± 0.2a B 0.98 ± 0.1a A 1.6 ± 0.1a B 3.4 ± 0.1a B 1.9a B 2.8b B

pH

T1 T2 T3 T4 T5 T6

6.5 ± 0.1 5.9 ± 0.2 5.8 ± 0.1 5.7 ± 0.2 5.5 ± 0.3 5.9 ± 0.1

T1 T2 T3 T4 T5 T6

28.3 ± 1.7 b D 32.5 ± 1.2 b E 25.4 ± 1.5 b D 21.4 ± 1.0 b C 13.3 ± 0.2 b B 0.92 ± 0.02 b A

15.3 ± 1.2 a E 4.12 ± 1.11 a D 0.64 ± 0.03 a C 0.14 ± 0.02 a B 0.08 ± 0.01 a B 0.03 ± 0.01 a A

T1 T2 T3 T4 T5 T6

2.43 ± 0.63 3.35 ± 0.07 2.11 ± 0.08 1.72 ± 0.09 0.79 ± 0.05 0.49 ± 0.02

aD bE bD bC bB bA

2.47 ± 0.06 2.14 ± 0.04 0.36 ± 0.02 0.28 ± 0.02 0.19 ± 0.02 0.11 ± 0.01

a a a a a a

T1 T2 T3 T4 T5 T6

2.60 ± 0.07 2.25 ± 0.02 1.36 ± 0.03 1.18 ± 0.03 1.31 ± 0.05 0.41 ± 0.00

b b b b b b

1.68 ± 0.08 1.09 ± 0.04 0.21 ± 0.01 0.07 ± 0.02 0.06 ± 0.02 0.06 ± 0.02

aD bC aB aA aA aA

T1 T2 T3 T4 T5 T6

24.4 ± 1.3 26.3 ± 9.4 22.4 ± 0.3 17.6 ± 0.9 10.9 ± 0.1 10.2 ± 0.1

T1 T2 T3 T4 T5 T6

31.2 ± 0.2 b D 35.3 ± 0.2 b E 29.2 ± 0.2 b C 24. 7 ± 0.4 b C 9.97 ± 0.34 b B 7.90 ± 0.94 b A

15.9 ± 0.1 c D 11.7 ± 0.0 b C 3.63 ± 0.01 a B 1.40 ± 0.14 a A 1.25 ± 0.12 a A 1.19 ± 0.05 a A

T1 T2 T3 T4 T5 T6

16.5 ± 0.5 b D 32.1 ± 0.7 b E 15.2 ± 0.5 b D 10.4 ± 0.3 b C 7.49 ± 0.12 b B 5.51 ± 0.12 b A

6.30 ± 0.21 a E 11.4 ± 0.14 a F 3.52 ± 0.1 a D 1.55 ± 0.13 a C 1.32 ± 0.09 a B 1.18 ± 0.046 a A

T1 T2 T3 T4 T5 T6

73.1 ± 0.3 92.2 ± 2.8 74.8 ± 1.3 62.1 ± 0.9 25.1 ± 0.8 19.8 ± 2.4

52.5 ± 3.6 a D 37.8 ± 1.4 a C 9.14 ± 0.24 a B 3.52 ± 0.35 a A 3.27 ± 0.29 a A 3.01 ± 0.12 a A

T1 T2 T3 T4 T5 T6

36.0 ± 11.4 b C 35.3 ± 4.2 b C 33.1 ± 3.5 b C 28.2 ± 2.8 b C 10. 8 ± 2.1 b B 4.61 ± 0.66 b A

17.1 ± 2.7 a D 10.1 ± 0.3 a C 1.87 ± 0.09 a B 0.48 ± 0.20 a A 0.55 ± 0.12 a A 0.57 ± 0.14a A

T0d T2 T4 T6

1.2 ± 0.2a A 1.1 ± 0.1a A 1.4 ± 0.2b B 1.6 ± 0.4b B

1.1 ± 0.1a 1.2 ± 0.3a 1.0 ± 0.0a 0.9 ± 0.0a

VFA (g CH3COOH l

NTK (g l

1

)

1

)

1

N–NH3 (g N l

)

Total alkalinity (g CaCO3 l

TOC (g l

1

DOC (g l

1

)

)

COD (g O2 l

BOD5 (g O2 l

1

)

1

)

WHCc (kg water kg dm

1

)

1

)

aa Bb aA aA aA aA aA

b b b b b b

b b b b b b

E D C B C A

B B B B A A

D E D C B A

6.7 ± 0.1 8.1 ± 0.3 9.0 ± 0.1 8.2 ± 0.4 8.3 ± 0.2 8.7 ± 0.3

b b b b b b

A B C B B B

F E D C B A

13.8 ± 0.9 a B 9.4 ± 0.3 a B 2.18 ± 0.09 a A 1.92 ± 0.05 a A 1.85 ± 0.06 a A 1.71 ± 0.05 a A

A A A A

a Means followed in the same line by the same lower-case letter are not statistically different (p < 0.05) according to Tukey test. b Means followed in the same column by the same capital letter are not statistically different (p < 0.05) according to Tukey test. c Water holding capacity of wastes measured during USMSWs incubation. d T0 = WHC measured on wastes at the start of the anaerobic incubation.

in simulated landfill, the S-USMSW leachates showed a TOC content that was about half of that of I-USMSW (Table 2). Moreover the TOC-leachate for stabilized waste (S-USMSW) decreased strongly during first 6 months of incubation to remain almost similar for the last part of the incubation. On the contrary, TOC-leachate of IUSMSW remained almost stable during all incubation until 8th month. At the end of the trial, TOC-leachate content in S-USMSW was 85% less than TOC-leachate of I-USMSW. As expected DOC, COD and BOD5 trends in leachates were similar to that of TOCleachate such as the good correlations found between these parameters considering both I-USMSW and S-USMSW data (n = 12) confirmed (TOC vs DOC: r = 0.89, p < 0.05, n = 12; TOC vs BOD5: r = 0.99, p < 0.05, n = 12; TOC vs COD: r = 0.99, p < 0.05, n = 12). Interesting was the fact that COD and BOD5 measured in the SUSMSW leachate at the end of the incubation period (COD of 3 g O2 l 1 and BOD5 of 0.57 g O2 l 1) were typically those of an old landfill (COD of 0.5–4.5 g O2 l 1 and BOD5 < 4 g O2 l 1) (El-Fadel et al., 1997; Kulikowska and Klimiuk, 2008). On the other hand, final COD and BOD5 values measured for I-USMSW leachate (COD of 19.8 g O2 l 1 and BOD5 of 4.61 g O2 l 1) were reported to be typically those l of young landfill (COD of 6–60 g O2 l 1 and BOD5 of 4–30 g O2 l 1) (El-Fadel et al., 1997; Kulikowska and Klimiuk, 2008). 3.2.4. Metal and anions content in leachates Investigation on leachates composition was completed with metals and anions determination. USMSW metal contents reported in Table 1, suggested that biological process led to a relative concentration of metal because of OM degradation. Heavy metal contents in leachates represented only a small fraction of the total HM contents (Table 3), in agreement with literature (Tatsi and Zouboulis, 2002). Data reported in Table 3 indicated different behaviors for the two wastes studied. I-USMSW showed a heavy metal reduction in leachates between sample T1 and sample T6 for Ni, Cd, Cu, Zn and Pb. Contrarily Hg, Cr, Ba and Sb contents after 1 year of incubation (sample T6), did not show any statistical difference with respect to starting samples (sample T1) (Table 3). In the last, Se, Mo and As contents increased during incubation. On the contrary, for the S-USMSW, the majority of metal contents decreased during incubation in simulated landfill, although Hg and Se remained unchanged and Mo content increased with respect to the initial content (sample T1). Total metals (lg l 1) leached during 1 years of incubation (Table 3) showed different behavior for the two theses studied; indeed Ni, Se, Mo and Zn concentrations in the leachates were higher in I-USMSW than in S-USMSW sample. On the contrary, Cu, Cr and Pb contents in leachates were higher in the S-USMSW than the I-USMSW, while Ni, Cd, Hg, Ba, As and Sb concentration were similar for the two wastes studied. Anion concentrations were always (F ) and often (CN ) below the detection limit of 1 mg l 1. Chlorides (Cl ) concentrations and its behavior was similar for both wastes, i.e. Cl decreased constantly until the end of incubation (Chu et al., 1994). At the end of the trials, Cl concentrations were lower than starting waste samples (Table 3). S-USMSW showed a concentration of SO42 lower than that for I-USMSW (Table 2) probably because of lower red-ox potential registered ( 103 and 160 mV for I-UFMSW and S-UFMSW, respectively), but, also, because of the presence of fresh organic matter rich in protein-like material containing sulfur. 3.3. The DRI to describe leachate impact reduction after MBT process The DRI have been extensively used in the past as biological stability index (Adani et al., 2004; Barrena et al., 2011; Sánchez-Arias

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S. Salati et al. / Bioresource Technology 128 (2013) 351–358 Table 3 Metal and anion content in leachates.

Ni

lg l

1

Time

I-USMSW

S-USMSW

T1

912 ± 94 aaBb 1.89%c 1232 ± 22 b C 1.01% 310 ± 31 a A 0.47% 277 ± 41 b A 0.94% 280 ± 24 b A 0.57% 290 ± 18 b A 0.37% 155 ± 34 a D 3.53% 68 ± 8 a C 0.61% 36 ± 3 b B 0.60% 8.3 ± 2.1 a A 0.30% 6.9 ± 3.2 a A 0.16% 10 ± 1 a A 0.14% 102 ± 2 a B 0.21% 71 ± 1 a A 0.06% 69 ± 21 a A 0.10% 167 ± 72 a B 0.55% 89 ± 12 a A 0.18% 67 ± 9 a A 0.08% 10 ± 2 a A 0.07% 21 ± 5 a B 0.06% 10 ± 1 b A 0.54% 4.9 ± 1.1 a A 0.06% 6.3 ± 1.8 a A 0.04% 9.4 ± 4.1 a A 0.04% 15 ± 3 b B 4.07% 4.4 ± 1.7 a A 0.43% 30 ± 1 a C 6.00% 44 ± 8 b C 19.42% 50 ± 7 b C 13.25% 58 ± 4 c C 9.79% 39 ± 20 a A 0.29% 71 ± 23 b A 0.21% 80 ± 3 a A 0.43% 146 ± 10 b B 1.74% 230 ± 1 C 1.64% 592 ± 19 b D 2.69% 8933 ± 48 b D 1.83% 8235 ± 268 b D 0.67%

1262 ± 82 b C 4.03% 464 ± 5 a A 0.66% 544 ± 60 b B 1.29% 72 ± 16 a A 0.37% 74 ± 2 a A 0.21% 89 ± 14 a A 0.37% 172 ± 37 a C 2.79% 35 ± 6 a B 0.25% 5.3 ± 1.1a A 0.06% 14 ± 2 a A 0.36% 10 ± 1 a A 0.14% 8.4 ± 3.3 a A 0.17% 1014 ± 10 b C 0.46% 829 ± 5 b B 0.17% 577 ± 109 b A 0.19% 503 ± 72 b A 0.36% 436 ± 51 b A 0.18% 396 ± 80 b A 0.23% 22 ± 5 b A 0.29% 33 ± 3 b A 0.20% 35 ± 6 a A 0.35% 12 ± 6 a A 0.26% 30 ± 3 b A 0.35% 32 ± 10 b A 0.55% 8.3 ± 0.9 a A 6.29 11 ± 1 b A 3.86% 24 ± 13 a A 14.00% 7.4 ± 1.2 a A 8.75% 7.1 ± 1.9 a A 4.86% 9.2 ± 2.9 a A 9.07% 25 ± 14 a A 0.96% 27 ± 4 a A 0.46% 231 ± 30 b B 6.58% 48 ± 17 a AB 2.93% 90 ± 17 a B 3.05% 103 ± 74 a B 5.07% 3146 ± 38 a C 1.23% 2678 ± 9 a B 0.47%

T2 T3 T4 T5 T6 Cd

T1 T2 T3 T4 T5 T6

Cu

T1 T2 T3 T4 T5 T6

Hg

T1 T2 T3 T4 T5 T6

Se

T1 T2 T3 T4 T5 T6

Mo

T1 T2 T3 T4 T5 T6

Zn

Table 3 (continued)

T1 T2

Time

I-USMSW

S-USMSW

T3

3549 ± 102 b C 0.54% 1452 ± 137 b B 0.48% 856 ± 15 a A 0.17% 978 ± 0.039 b A 0.12% 62 ± 7 a A 0.11% 54 ± 7 a A 0.04% 46 ± 5 a A 0.61% 115 ± 12 a B 0.34% 65 ± 11 a A 0.11% 39 ± 3 a A 0.04% 251 ± 1 a C 0.21% 158 ± 3 a B 0.05% 150 ± 31a B 0.09% 564 ± 7 b D 0.78% 122 ± 11 a B 0.10% 44 ± 10 a A 0.02% 393 ± 37 a A 0.09% 369 ± 117 a A 0.04% 350 ± 93 a A 0.06% 615 ± 4 b C 0.24% 521 ± 3 b B 0.12% 304 ± 16 b A 0.05% 40 ± 3 a A 1.15% 135 ± 31a B 1.54% 141 ± 25 a B 2.99% 154 ± 44 a B 7.21% 110 ± 21 a B 3.09% 90 ± 37 a B 1.61% 131 ± 25 a A 9.23% 55 ± 17 a A 1.54% 58 ± 12 a A 3.01% 57 ± 10 a A 6.53% 68 ± 12 a A 4.68% 123 ± 18 b A 5.39% 2.00 ± 0.02 a C 1.69 ± 0.34 a C 1.45 ± 0.11 a C 1.39 ± 0.08 b C 0.64 ± 0.02 a B 0.34 ± 0.04 a A <1.00

1346 ± 44 a A 0.39% 1221 ± 28 a A 0.76% 1232 ± 35 b A 0.43% 1223 ± 125 a A 0.61% 969 ± 51 b F 2.04% 581 ± 33 b D 0.55% 722 ± 82 b E 1.13% 142 ± 1 b C 0.48% 76 ± 3 b B 1.41% 39 ± 6 b A 0.11% 1670 ± 3 b E 0.37% 1710 ± 21 b E 0.17% 695 ± 30 b D 0.11% 293 ± 6 a C 0.10% 197 ± 2 b B 0.04% 104 ± 14 b A 0.03% 856 ± 38 b E 0.38% 647 ± 7 b D 0.13% 243 ± 56 b A 0.08% 374 ± 3 a C 0.27% 301 ± 5 a B 0.12% 268 ± 9 a A 0.15% 141 ± 40 b B 2.30% 165 ± 4 b B 1.20% 290 ± 20 b C 3.52% 115 ± 12 b AB 2.99% 103 ± 11 a A 1.49% 94 ± 2 a A 1.97% 136 ± 2 a B 7.75% 97 ± 4 b A 2.47% 87 ± 10 b A 3.68% 85 ± 12 b A 7.70% 81 ± 7 a A 4.08% 72 ± 6 a A 5.27% 2.64 ± 0.10 b E 2.87 ± 0.21 b E 1.57 ± 0.09 a D 1.02 ± 0.06 a C 0.56 ± 0.04 a B 0.39 ± 0.01 a A <1.00

T4 T5 T6 Cr

T1 T2 T3 T4 T5 T6

Pb

T1 T2 T3 T4 T5 T6

Ba

T1 T2 T3 T4 T5 T6

As

T1 T2 T3 T4 T5 T6

Sb

T1 T2 T3 T4 T5 T6 1

Chlorides

gl

Fluorides

mg l

1

T1 T2 T3 T4 T5 T6 T1

(continued on next page)

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Table 3 (continued)

1

Sulfates

gl

Cyanides

mg l

1

Time

I-USMSW

S-USMSW

T2 T3 T4 T5 T6 T1 T2 T3 T4 T5 T6 T1 T2 T3 T4 T5 T6

<1.00 <1.00 <1.00 <1.00 <1.00 0.79 ± 0.09 0.67 ± 0.11 0.80 ± 0.05 1.33 ± 0.10 0.97 ± 0.07 0.55 ± 0.04 <0.01 <0.01 <0.01 <0.01 <0.01 0.03

<1.00 <1.00 <1.00 <1.00 <1.00 0.37 ± 0.03 1.12 ± 0.08 0.17 ± 0.02 0.18 ± 0.03 0.21 ± 0.03 0.22 ± 0.05 <0.01 <0.01 0.01 0.04 0.02 0.02

bA aA bA bB bB bA

aB bC aA aA aA aA

a Means followed in the same line by the same lower-case letter are not statistically different (p < 0.05) according to Tukey test. b Means followed in the same column by the same capital letter are not statistically different (p < 0.05) according to Tukey test. c Percentage value referred to absolute value of metal content of wastes determined by mass balance metal value determined by mass balanced (mg per column).

et al., 2012) to predict waste impacts i.e. biogas and odors productions coming from biomass, COD and BOD5 of waste-leachates (Scaglia and Adani, 2008) and self-heating of the biomass, since they well correlated with DRI (Scaglia et al., 2011). Italian law (Gazzetta Ufficiale, 2010) indicates a DRI value below 1000 mg O2 kg VS 1 h 1 for biological treated USMSW to be landfilled, as this limit should guarantee a strong reduction of potential impacts of waste in landfill.

The results of this work indicated that 28 days of full scale aerobic biological treatment of USMSW (I-USMSW) determined a degradation of OM contained in the waste getting a high degree of biological stability, i.e. DRI = 978 mg O2 kg SV 1 h 1. In order to understand the real impact of the leachate of pre-treated material in landfill with respect to that of untreated waste, both no-biostabilized (I-USMSW) and biostabilized (S-USMSW) wastes were incubated for 1 year in a simulated landfill and leachate characteristics compared. Results obtained (Table 2) indicated that biological treatment greatly altered the successive anaerobic biological processes occurred in the simulated landfill giving different leachate for untreated and treated waste. Low pH and, high TKN, NH3, and VFA acid contents characterizing leachate of untreated waste (I-USMSW) at the end of the incubation (Table 2), were those typical of a young landfill characterized by an acid phase. This phase is characterized by an intense hydrolysis of fresh organic matter and fermentation of hydrolyzed monomers producing, mainly, organic acids (VFA) (Kjeldsen et al., 2002) that were, in effect, well represented in leachate (Table 2). This fact was confirmed by the very good positive correlation found for TOC vs. VFA contents in the leachate (r = 0.96; p < 0.05; n = 6). High VFA and low pH (Table 2) were responsible for the partial inhibition of the biological process occurred in the simulated landfill of untreated waste, slowing down degradation processes (Kjeldsen et al., 2002; Tatsi and Zouboulis, 2002) during the first 8 months (sample T4) of incubation (Table 2). Degradation process became considerable only in the last stage of the incubation (8– 12th month) (samples T4–T6), when inhibitory condition were partially removed (Table 2). This fact was confirmed by the strong reduction of both TOC and VFA contents in the leachate in the last stage of incubation (see Table 2).

Fig. 1. Total TOC, TKN, NH3, COD and BOD5 leached (g kg

1

dm) for I-USMSW and S-USMSW.

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On the other hand, leachate coming from biostabilized waste (SUSFMSW) showed different trend. High pH, and low TKN, N–NH3 and VFA contents, indicated an advanced stage of OM decomposition (Kjeldsen et al., 2002). Under this state inhibitory conditions were avoided and although all parameters measured at the start of the incubation (TOC, DOC, NTK, NH3, DOC and BOD5) were all already lower than those indicated for the untreated waste (Table 2), degradation process proceeded fast and all parameters further strongly decreased before the end of incubation. After just 6 months of incubation leachate did not show any appreciable variation of its characteristics, except for BOD5 that continued to decrease till the end of the incubation. Therefore, bio-stabilization process allowed the degradation of easily degradable OM reducing impact of leachate in simulated landfill and, in addition, optimal condition for degradation process were guaranteed allowing just after 1 year of incubation in simulated landfill getting leachate characteristics typical of an old landfill (Kjeldsen et al., 2002). Contrarily, untreated waste contributed to a more polluted leachate and to the partial inhibition of biological process, delaying OM stabilization in the simulated landfill. As consequence of that at the end of incubation, of leachate that was typical of that of a young landfill under acidic phase (Kjeldsen et al., 2002). If these results were expected and already discussed in literature (Kjeldsen et al., 2002; Tatsi and Zouboulis, 2002), less is known about the total reduction of the potential impacts of leachate of a biostabilized material characterized by a DRI below 1000 mg O2 kg SV 1 h 1. Taking into consideration total leachate production and data detected for parameters at the end of the incubation period (Table 2), it was calculated the total reduction of the impact of leachate for biostabilized vs. un-biostabilized incubated waste, i.e. 54%, 69%, 77%, 70% and 81% for NTK, NH3, TOC, COD and BOD5, respectively (Fig. 1). These data seem to confirm that the achievement of biological stability after only 28 d of full scale biological process, allowed getting a strong reduction of leachate impacts in landfill. Besides less clear is the effect of waste pre-treatment on inorganic chemical species leaching, i.e. heavy metals. The total amounts of heavy metals leached from biostabilized and un-biostabilized wastes incubated in the simulated landfills were considered in this work by calculating the total moles of HM leached. This was made because the use of total weight (g)

gives erroneous interpretation of the results since each metal is characterized by different atomic weight (Fig. 2). Biological process because of degradation of organic matter determined a concentration of total heavy metal contents that was quantified for S-USMSW, by taking into consideration total leachate production of I-USMSW and S-USMSW and the relative HM contents (as mol l 1), to be 7% higher than HMs content of IUSMSW. Despite to the HM higher content than untreated waste, after 1 year of incubation in the simulated landfill, total HM (moles) leached from biostabilized material was 16% lower than that of no-biostabilized material (Fig. 2a). Therefore it can be concluded that biostabilization process limited HM leaching in successive waste landfilling. Heavy metal immobilization in landfill depends both by sorption phenomena of HMs on organic matter and mineral and, above all, by precipitation reaction because of the presence of carbonate, sulfides and hydroxide anions (Kjeldsen et al., 2002). The higher leachability of HMs for I-USMSW than S-USMSW, depended on the presence of low pH that directly influenced mobilization/immobilization of HMs (Kjeldsen et al., 2002), i.e. low pH did not favor both hydroxide formation and availability of carbonate anions forming insoluble HM salts. In addition the high presence of soluble OM measured in the leachate for untreated waste acting as organic ligands, probably, mobilized HMs (see in Table 2 TOC, DOC and VFA contents) (Kjeldsen et al., 2002; Bilgili et al., 2007). On the opposite high pH and low presence of soluble OM in S-USMSW did not favor HMs mobility. Good correlation found for leached TOC vs. HMs (as moles) (r = 0.69, p < 0.05, n = 12) confirmed this hypothesis. Nevertheless, when mass balances were performed for each HM leached, data showed contrasting results for some metals, i.e. Cu, Cr and Pb concentration in the leachates were higher for S-USMSW than I-USMSW. These data appeared in contrast with what above discussed additionally that at pH 7– 8, (pH measured for S-USMSW), both Pb and Cu are typically present under hydroxide forms. An explanation of these results can be found taking into consideration that Cu and Pb can form sulfide complex characterized by very low solubility in water (solubility constant, ksp, of 8.5  10 45 and of 3  10 28 for Cu and Pb, respectively) (Xiaoli et al., 2007). Sulfides originated by sulfates reduction during waste decomposition in landfill, and sulfide insoluble forms were often cited as root of low mobility of heavy metals in landfill (Kjeldsen

Fig. 2. Total heavy metal leached for I-USMSW and S-USMSW, (a) as mmol kg

1

dm; (b) as mg kg

1

dm.

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et al., 2002). In effect, the incubation of the I-USMSW led to the presence of high amounts of sulfide as consequence of anaerobic degradation of fresh OM rich in sulfur-containing molecules (e.g. proteins). Unfortunately, no direct measurement of these chemical specie was made; nevertheless, higher sulfate contents in IUSMSW than S-USMSW, seem to confirm this hypothesis, i.e. 11 and 3.81 g of sulfate leached from I-USMSW and S-USMSW respectively. The behavior of Cr is not easy to be explained. The main Cr species under reduced condition are hydroxide complexes (Baun and Christensen, 2004; Kjeldsen et al., 2002) that show a very low water solubility (KspCr(OH)3 of 6.3  10 31). Optimal conditions for hydroxide complex formation are high pH that was typical for S-USMSW and not for I-USMSW, and so in contrast with the results before proposed. Moreover, literature (Xiaoli et al., 2007) suggested that Cr might be strongly bind to hydroxides, precipitate as sulfides or absorbed on organic substances that should determine its lower mobility in I-USMSW rather than S-USMSW. Nevertheless, Kim et al. (2011) observed a different behavior of Cr in leachates of anaerobic bioreactors respect with respect to other metals; indeed, while metals showed the classic amphoteric shape with the greatest leachability in a pH-range of 4.5–7.0, Cr showed higher leaching value in a pH-range 7.0–8.5. 4. Conclusion The DRI can be used to predict potential impact of landfilled waste with particular reference to leachate; nevertheless, no data existed regarding direct measurement of these impact reductions. Full scale waste MBT treatment and successive incubation in simulated landfill for 1 year gave the following leachate impacts reduction: 54%, 69%, 77%, 70%, 81% and 16% for NTK, NH3, TOC, COD, BOD5 and total heavy metal contents, respectively. These results indicated that short MBT process getting a DRI < 1000 mg O2 kg VS 1 h 1 allows a strong reduction of leachate pollution. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2012.10. 101. References Adani, F., Confalonieri, R., Tambone, F., 2004. Dynamic respiration index as descriptor of the biological stability of organic wastes. J. Environ. Qual. 33, 1866–1876. Allen, A., 2001. Containment landfills: the myth of sustainability. Eng. Geol. 6, 3–19. APHA-American Public Health Association, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. APHA, Washington DC, USA. Barrena, R., Turet, J., Busquets, A., Farrés, M., Font, X., Sánchez, A., 2011. Respirometric screening of several types of manure and mixtures intended for composting. Bioresour. Technol. 102, 1367–1377. Barlaz, M.A., 2006. Forest products decomposition in municipal solid waste landfills. Waste Manage. Res. 26, 321–333. Baun, D.L., Christensen, T.H., 2004. Speciation of heavy metals in landfill leachate: a review. Waste Manage. Res. 22, 3–23.

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