Mechanical biological treatment of organic fraction of MSW affected dissolved organic matter evolution in simulated landfill

Bioresource Technology 142 (2013) 115–120

Contents lists available at SciVerse ScienceDirect

Bioresource Technology journal homepage: www.elsevier.com/locate/biortech

Mechanical biological treatment of organic fraction of MSW affected dissolved organic matter evolution in simulated landfill 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 Srl, Italy

h i g h l i g h t s  Dissolved organic matter (DOM) controls many processes in landfill.  Aerobic pretreatment affects successive DOM evolution in landfill.  DOM evolution of treated and untreated organic waste in landfill was study.  Aerobic pretreatments reduce DOM because of stabilization of organic matter.  Untreated waste reduce DOM because biological process, hydrolysis, was stopped.

a r t i c l e

i n f o

Article history: Received 15 March 2013 Received in revised form 14 May 2013 Accepted 15 May 2013 Available online 21 May 2013 Keywords: Dissolved organic matter Hydrophilic fraction Hydrophobic fraction Landfill Mechanical biological treatment

a b s t r a c t The aim of this paper was to study the evolution of DOM during 1 year of observation in simulated landfill, of aerobically treated vs. untreated organic fraction of MSW. Results obtained indicated that aerobic treatment of organic fraction of MSW permitted getting good biological stability so that, successive incubation under anaerobic condition in landfill allowed biological process to continue getting a strong reduction of soluble organic matter (DOM) that showed, also, an aromatic character. Incubation of untreated waste gave similar trend, but in this case DOM decreasing was only apparent as inhibition of biological process in landfill did not allow replacing degraded/leached DOM with new material coming from hydrolysis of fresh OM. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Landfill represents the most used method for municipal solid wastes (MSW) disposal in the world because it requires simple operations and it represents the cheapest method for waste disposal (Allen, 2001). On the other hand MSW landfilling causes impacts such as the production of biogas, odors and leachate because of the presence in waste of biodegradable organic matter (OM) (Scaglia et al., 2010). In order to reduce waste impacts in landfill, mechanical biological treatment (MBT) of MSW has been proposed in the past and now it represents a diffuse waste pre-treatment before waste landfilling (Adani et al., 2004; Lornage et al., 2007). MBT before landfilling allows reducing organic matter contained in the waste with particular reference to the more degradable organic fraction because of aerobic biological process, reducing potential

⇑ Corresponding authors. Tel.: +39 0250316544; fax: +39 0250316521. E-mail addresses: [email protected] (B. Scaglia), [email protected] (F. Adani). 0960-8524/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2013.05.049

waste impacts (European Parlament, 1991; El-Fadel et al., 2002; Scaglia and Adani, 2008; Scaglia et al., 2010). Aerobic biological process determined, also, qualitative modification of OM. Typically OM degradation proceeds by hydrolysis of polymer (e.g. cellulose, proteins, lipids and hemicelluloses) to dissolved organic matter (DOM) e.g. simple sugar, aminoacid and fatty acid, and the successive its degradation by microorganism (Said-Pullicino et al., 2007). In this way aerobic process determines the degradation of easily degradable organic fraction and the concentration of more recalcitrant ones, i.e. lignin derived molecules and recalcitrant lipids (Adani et al., 1995; De Gioannis et al., 2009). Fresh organic matter is characterized by high amount of DOM contrarily to degraded/stabilized OM (Said-Pullicino et al., 2007). So DOM represents a parameter related to the acquirement of the biological stability such as previously reported (Iannotti et al., 1993; D’Imporzano and Adani, 2007). From a quantitative point of view DOM represents only a small part of the organic matter of waste; nevertheless it is considered the most mobile and reactive organic fraction controlling physical,

116

S. Salati et al. / Bioresource Technology 142 (2013) 115–120

chemical and biological processes in landfill. DOM is able forming chemical and/or physical interaction with both metals and organic xenobiotics (Zsolnay, 1996). Christensen and Christensen (1999) observed an increase in metal contents bound to DOM when concentration of this fraction increased in leachated-polluted groundwater. DOM consists of both hydrophilic (organic acids, carbohydrates, amino acids and amino sugars) and hydrophobic (aromatic phenols, hydrocarbons, fats, and nucleic acids) components (Dilling and Kaiser 2002; Said-Pullicino et al., 2007; Zi-gang et al., 2007), although biological process affects DOM characteristics (D’Imporzano and Adani, 2007). DOM extracted from fresh organic matter is characterized by a large number of low MW hydrophilic and hydrophobic molecules. Organic matter degradation determined a decrease of both hydrophilic and hydrophobic fractions leading to the concentration of more recalcitrant ones (aromatic-like molecules) (D’Imporzano and Adani, 2007), and to different behavior of DOM vs. metal and/or organic pollutants (Zi-gang et al., 2007). No many data exist in literature about the fate of DOM in landfill. In addition, no data exist regarding the effect of aerobic pretreatment on the fate of DOM in landfill. Being MBT process widely diffuse in EU and now, also, in many other countries, it becomes of interest to understand the effect of MBT process on DOM content in waste taking into consideration that this parameters plays a central role in the mobilization of inorganic and organic pollutants. The aim of this paper was to study the evolution of DOM during one-year observation in simulated landfill of aerobically treated vs. untreated organic fraction of MSW coming directly from a fullscale MBT plant.

2. Methods 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 d1 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) (Salati et al., 2013). Operating temperature of bio-stabilization was maintained in the range of 45–50 °C, 50–65 °C and 50–55 °C during 1st, 2nd3rd and 4th week respectively, by regulating forced aeration.

2.2. MSW samples The USMSW was sampled at the start (I-USMSW, undersize fraction Ø < 90 mm of MSW) and after 28 d of the biological process (S-USMSW). Yellow flags were used to mark off waste sections and to find S-USMSW after 28 days of treatment. Samples were taken such as previously reported (Scaglia et al., 2011) by using standard sampling procedures (European Committee for Standardization, 2006) assuring representativeness of the samples. In brief, starting from a waste mass of about 6000– 7000 kg, new waste pile of about 2000 kg were obtained by a series of successive weight reductions (CEN, 2006). Then, from this pile, a final sample of 40–50 kg w.w., was obtained by collecting waste increments. Each increment was sampled by applying the procedure of the stratified random sampling (CEN, 2006). The increased size of each waste was calculated by considering the size of the waste particle and the heterogeneity degree of the MSW (CEN, 2008). The sample was then brought to the laboratories, stored

at 4 °C, and processed within 3–5 days from the date of sample receiving. 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 1 kg of dry matter was reduced in size (particle size < 1 mm) and used for successive analysis and DOM extraction. 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) (Salati et al., 2013). 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 water hold capacity (WHC) of 75% w/w (Scaglia et al., 2010), resulting a final bulk density of the mass of 0.8 Mg m3. 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 (Salati et al., 2013). 2.4. DOM extraction Dissolved organic matter (DOM) was extracted from waste samples at different time of incubation in the simulated landfills, i.e. 0, 4, 8 and 12 months of incubation. DOM was extracted by using method reported by D’Imporzano and Adani (2007); 5 g of dried material was extracted by water (1:20 solid:liquid ratio, w/w) using a Dubnoff bath at 60 rpm for 30 min at 40 °C. To obtain the highest DOM yield and to avoid hydrolysis, extraction parameters were those used previously (D’Imporzano and Adani, 2007). Afterwards the suspension obtained was centrifuged for 15 min at 6500 rpm and filtered twice: firstly by a fast cellulose filter (Whatman paper filter N.4) and then by a 0.45 lm Millipore membrane (Advantec MFS, Pleasanton, CA). Solution obtained represented the dissolved organic matter. 2.5. DOM fractionation DOM was fractionated into hydrophilic (Hi), hydrophobic (Ho), and neutral hydrophobic (NHo) fractions according to the procedure proposed by Leenheer (1981), that was partially modified. Two hundred ml of DOM, previously acidified at pH < 2 (0.5 mol l1 of H2SO4) were loaded onto glass column loaded with Amberlite XAD-7 (Sigma–Aldrich Steinheim, Germany) (40 cm high, 2.5 cm of diameter) previously activated with 0.5 mol l1 of H2SO4 and then washed until neutral pH and eluted at the velocity of 2 ml min1. Then, the column was washed with distilled water (2 bed volume); the eluted fraction represented the Hi fraction. After that the Ho fraction was eluted by using 0.05 mol l1 of NaOH (1.25 bed volume) and distilled water (2 bed volume). Ho fraction was then cleaned from Na+ by using a cation exchange resin (Amberlite IR 120, Merck, Darmstad, Germany). DOM fractionation was performed in triplicate; part of fractions obtained was immediately used for biodegradability test and part dried at 60 °C under vacuum for successive analyses. All fractions, i.e. DOM, Hi, Ho, were quantified by measuring total carbon content (APHA, 1998) obtaining DOM-C, Hi-C, Ho-C. NHo-C was calculated by subtracting the sum of Hi-C and Ho-C from the DOM-C.

117

S. Salati et al. / Bioresource Technology 142 (2013) 115–120

2.6. Degradability tests

Table 1 MSW characterization.

Solid matrices and water soluble fractions, i.e. DOM, Hi and Ho, obtained at different incubation time underwent to biodegradability test, which measured the oxygen uptake rate under liquid condition (OURL) to degrade easily degradable organic matter (D’Imporzano and Adani, 2007). In brief, oxygen uptake rate was measured for 20 h under liquid condition using 100 ml of extract (Scaglia and Adani, 2009). During the test, standard conditions were maintained so that optimal microbial activity and reaction rates were obtained. In particular, so as to keep optimal pH and to guarantee available nutrient, 5 g of dry matter of USMSW (or aqueous extracts obtained from 5 g of waste dry matter) were set in a flask and added of 500 ml of deionized water, 12 ml of phosphate buffer solution (KH2PO4 0.062 mol l1, K2HPO4 0.125 mol l1, Na2HPO47H2O 0.125 mol l1, pH 7.2), and 5 ml of nutritive solution (CaCl2 0.25 mol l1, FeCl3 0.9 mmol l1 and MgSO4 0.09 mol l1 made up according to the standard BOD test method procedures (APHA, 1998). No nitrogen was added to the solution, except for Ho fractions, as previous tests revealed that N content limited degradation of this fraction. In order to allow oxygen to correctly diffuse into suspension, the liquid extracts were agitated and aerated every 15 min. A feedback control system based on O2 measurement allowed optimal oxygen concentration (O2 > 2 mg l1). OURL was detected by measuring the slope of the decrease of the oxygen concentration in the matrices and the liquid extracts (DOM, Hi and Ho fractions) during the absence of aeration (D’Imporzano and Adani, 2007). Biodegradability test were conducted at 35 ± 2 °C. The cumulative oxygen consumption during the 20 h test-length (OD20) represented the degree of biodegradability of dissolved soluble fractions based on carbon content of each fraction. The OD20 were calculated using the following equation (Eq. (1)):

OD20 ¼

V  mðd:m:=1000Þ  ðC=1000Þ

Z

t¼20

jSjtdt

ð1Þ

I-USMSW d.m. VS TKN DRI

-1

(g kg w.w. ) (g kg d.m.-1) (mg O2 kg VS1 h1)

44.4 ± 1.56a 556 ± 19a 11.7 ± 0.8a 3910 ± 358b

S-USMSW *

47.5 ± 2.03a 525 ± 15a 11.3 ± 1.9a 978 ± 153a

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

3. Results and dicussion 3.1. Biological process MBT process determined a degradation of organic matter contained in the waste, i.e. TOC and VS contents reduced during the process by about 30 g kg1 on a relative basis (Table 1). NTK relative (g kg1 d.m.) content did not show significant reduction during biological process depending by the low degradation of the organic matter (Table 1). Degradation leaded to stabilization of waste such as indicated by DRI values (Table 1) that after 28 d of biological process (SUSMSW) was of 978 ± 153 mg O2 kg VS1 h1 (Tambone et al., 2001; Salati et al., 2013) which fully respected Italian rules for landfill disposal of biological pre-treated waste (DRI < 1000 mg O2 kg VS1 h1) (Gazzetta Ufficiale, 2010). Data fully respect the real potential degradability of wastes as analyses were performed on fresh sample just arrived in the laboratory (Pognani et al., 2012). Chemical parameters characterizing untreated (I-USMSW) and MBT-treated USMSW (S-USMSW) are resumed in Table 1.

3.2. Dissolve organic matter behavior during incubation in simulate landfills

t¼0

Where V is the volume of the suspension (l), m is the mass of the sample (g w.w.), d.m. is the dry matter content, C is the dissolved carbon content of each fraction (DOM-C, Hi-C, Ho-C for DOM, Hi, and Ho respectively) and |S|t is the rate of oxygen consumption at time t (mg O2 l1 min1).

2.7. Spectroscopic analysis Compost fractions were analyzed by Diffuse Reflectance Infrared Fourier Transformed (DRIFT) spectroscopy using an Avatar 370 FTIR from Thermo Nicolet Instruments (Madison, WI, USA). Samples (7 mg), previously dried at 65 °C for 48 h, and KBr (700 mg; FT grade, Aldrich Chemical Co, ST Louis, Missouri) were finely ground for 10 min using an agate ball mill (Specamill-Greseby-Specac, Kent, UK). Instrument parameters used were: scanning 128, resolution 4 cm1, and frequency 400–4000 cm1 and gain 16.

2.8. Statistical analysis Results obtained were analyzed by ANOVA, statistical difference between means was assessed using Tukey test. Chemical analyses were performed on three analytical samples taken from the 300 g composite bulk sample, and since standard deviation values were calculated using the data obtained from these three replicates, they represent estimates of the variability caused by both the waste bed homogeneity and the analytical method.

I-USMSW showed a DOM-C contents that was much higher than that of stabilized waste (S-USMSW) (Table 2). This data agreed with literature that reported as aerobic biological process allows reducing dissolved organic matter in stabilized material (D’Imporzano and Adani, 2007). The decrease of DOM-C was likely due to the reduction of Hi-C and Ho-C fractions (Table 3) which are generally considered to be more degradable than NHo-C fractions (Said-Pullicino et al., 2007; Shao et al., 2009). DOM-C of untreated waste was composed, above all, by Hi-C (65% of DOM-C) followed by NHo-C (20% of DOM-C) and Ho-C (16% of DOM-C). On the other hand DOM-C for treated waste was different as it was composed, above all, by the NHo-C fraction (46% of DOM-C) followed by an equal contribution of Hi-C and HoC (27% of DOM-C). These data agreed with previous reports about the effect of aerobic biological process on DOM-C fractions distribution (D’Imporzano and Adani, 2007), i.e. NHo-C fraction relatively increased after aerobic biological process. Landfill incubation determined different trend in the evolution of DOM-C: in the case of untreated waste (I-USMSW) DOM-C decreased rapidly during first 8 months of incubation to remain stable later (Table 3). On the other hand for the stabilized waste (S-USMSW), DOM-C decreased dramatically during first 4 months to remain practically constant for the rest of the incubation period. DOM-C contents after 12 months of incubation were low with respect to the starting values, although for the untreated waste, the content was double than that of treated waste, on a relative basis (g kg1 dm) (Table 3). On the other hand total DOM-C reduction, as absolute value (g), was quite similar for wastes studied and they were, taking into consideration mass balance and relative contents

118

S. Salati et al. / Bioresource Technology 142 (2013) 115–120

Table 2 Carbon content in DOM-C, Hi-C, Ho-C and NHo-C fractions during MBT and 12-months incubation period. Time (month)

0 4 8 12 * §

I-USMSW

S-USMSW #

DOM-C (g kg d.m.1)

Hi-C

Ho-C

NHo-C

28.46 ± 0.14c*C§ 11.03 ± 0.06bB 4.11 ± 0.08cA 4.60 ± 0.14bA

18.19 ± 0.84bC 3.77 ± 0.09bB 0.62 ± 0.09aA 0.62 ± 0bA

4.58 ± 1.21aB 1.28 ± 0.06aA 0.88 ± 0.03aA 0.95 ± 0.14bA

5.69 ± 1.21aAB 5.98 ± 0.08bB 2.61 ± 0.07bA 3.03 ± 0.02bA

DOM-C

Hi-C

Ho-C

NHo-C#

18.23 ± 0.64bC 3.49 ± 0.07aB 1.71 ± 0.05aA 2.19 ± 0.21aA

4.95 ± 0.06aC 1.61 ± 0.04aB 0.7 ± 0.04aA 0.57 ± 0.01bA

4.88 ± 0.18aD 1.76 ± 0.02bC 0.93 ± 0.19abB 0.19 ± 0.01bC

8.4 ± 0.78bC 0.12 ± 0.08aA 0.08 ± 0.06aA 1.43 ± 0.06aB

Means followed in the same line by the same lower-case letter are not statistically different (p < 0.05) according to Tukey test. Means followed in the same column by the same capital letter are not statistically different (p < 0.05, according to Tukey test). # Obtained subtracting to DOM-C the Hi-C and Ho-C fractions.

Table 3 OD20 values of DOM-C and its fractions referred to own C content. Sample I-USMSW

S-USMSW

a

Time (month) 0 4 8 12 0 4 8 12

DOM-C (mg O2 g1 DOM-C 20 h1) a

1138 ± 15 b 1460 ± 12c 2579 ± 4d 612 ± 46a 323 ± 8a 905 ± 63b 894 ± 29b 1023 ± 87b

Hi-C (mg O2 g1 Hi-C 20 h1)

Ho-C (mg O2 g1 Ho-C 20 h1)

1397 ± 56c 245 ± 43b 61 ± 11a 2361 ± 91d 1093 ± 18c 181 ± 16a 117 ± 10a 350 ± 30b

1378 ± 118c 460 ± 44a 829 ± 77b 2405 ± 102d 1076 ± 22c 504 ± 24a 789 ± 38b 976 ± 43d

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

of DOM-Cs, of 87% of initial DOM-C and of 89 % of initial DOM-C for I-USMSW and S-USMSW, respectively. During incubation, DOM-C reduction for untreated waste was due to both Hi-C and Ho-C decrease, that contributed proportionally to the DOM-C reduction. This fact was confirmed by the very good correlations found for DOM-C vs. Hi-C (r = 0.99, p < 0.05, n = 4) and DOM-C vs. Ho-C (r = 0.98, p < 0.05, n = 4). On the other hand the NHo-C fraction decreased less than the others and, at the end of the process, the DOM-C was formed for 65% of this fraction. S-USMSW showed a different trend as all these three DOM-C fractions, i.e. Hi-C, Ho-C and NHo-C, contributed to the total DOM-C reduction, such as the good correlations found confirmed (DOM-C vs. Hi-C: r = 0.99, p < 0.05, n = 4; DOM-C vs. Ho-C: r = 0.96, p < 0.05, n = 4; DOM-C vs. NHo-C: r = 0.98, p < 0.05, n = 4). Such as showed for the untreated waste at the end of the incubation the NHo-C fraction contributed mainly to DOM-C formation, i.e. 65% of DOC. As consequence of rain events, DOM-C was partially leached during incubation experiment (Salati et al., 2013). Leachate productions resulted similar for both wastes incubated (respectively of 11.6 l and 12.9 l for I-USMSW and S-USMSW respectively) however, S-USMSW DOM-C content was about half of that of I-USMSW (Table S1). Moreover, the DOM-C-leachate for S-USMSW decreased strongly during first 6 months of incubation to remain almost similar for the last part of the incubation. On the contrary, DOM-C leachate of I-USMSW remained almost stable during all incubation until 8th month. At the end of the trial, DOM-C leachate content in S-USMSW was 79% less than DOM-C leachate of I-USMSW (Salati et al., 2013). 3.3. Biodegradability of DOM and DOM-fractions Biodegradability test were assessed in order to get qualitative information about DOM and its faction. NHo degradability was not tested as this fraction was not extracted. DOM for the untreated waste (I-USMSW) was more degradable than those for treated waste (S-USMSW). DOM degradability

increased with the process and highest value was measured after 8 months of incubation reducing drastically after 12 months. This trend is not easy to be explained. Data reported in Table 3, represent a qualitative data, i.e. degradability of DOM-C (mg O2 g1 DOM-C 20 h1) and not quantitative data, i.e. total oxygen uptake due to DOM-C in waste (mg O2 g1 waste d.m. 20 h1). Therefore results seem to indicate that more oxidized molecules were present in the initial stage of anaerobic incubation, i.e. VFA (see Table 4). Successively VFA were degraded (Table 4) and substituted by more reduce molecules. This does not mean that oxygen uptake rate increased as absolute data; in fact OURL when referred to the dry matter waste unit (calculated starting from data of DOM-C in Tables 2 and 3) decreased, i.e. 32.3, 16.1, 10.6 and 2.81 mg O2 g1 d.m. 20 h1 after 0, 4, 8 and 12 months, respectively. On the other hand both Hi-C and Ho-C biodegradability trend was, unexpectedly, different as they decreased till 8th months to increase a lot at the end of incubation. Results indicated a contradiction between low degradability of DOM-C and high degradability of Hi and Ho. This paradox can be explained taking

Table 4 Characterisation of USMSWs during the landfill incubation phase.

C (g kg1d.m.)

pH

VFA (mg CH3COOH l1)

Time (month)

I-USMSW

S-USMSW

0 4 8 12 0 4 8 12 0 4 8 12

298 ± 15abbA 280 ± 9bA 271 ± 4bA 290 ± 3bA 5.6 ± 0.2aA 6.6 ± 0.3aB 6.5 ± 0.1aB 6.6 ± 0.2aB 4123 ± 21bD 700 ± 16bC 280 ± 4bB 88 ± 0bA

267 ± 8aC 207 ± 17aC 163 ± 3aA 183 ± 4aB 6.8 ± 0.2bA 7.6 ± 0.3bB 7.5 ± 0.3bB 7.8 ± 0.2bB 816 ± 197aD 119 ± 26aC 69 ± 6aB 17 ± 4aA

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.

S. Salati et al. / Bioresource Technology 142 (2013) 115–120

into consideration that i. at the end of the incubation these two fractions, quantitatively, contributed for less than half to DOM-C (Table 2), ii. that the NHo-C fraction contributed for 65% to DOMC (Table 2) and iii. that, this fraction had very low biodegradability (D’Imporzano and Adani, 2007). DOM degradability for treated waste showed a different trend as its content increased after just 2 month of incubation, remaining then constant till the end of the process. This trend can be explained taking into consideration that waste at time zero had low DOM-C content as it was coming from aerobic process during which soluble organic matter was readily degraded. Probably, successive incubation under anaerobic condition led to the increase of DOM-C that was, in any case, limited because of biological stability acquired during aerobic treatment. DOM degradability for S-UFMSW was constant during the process except for the first stage of incubation during which it was low. Such as before stated, at the start of the process DOM characteristic were affected greatly by aerobic process before performed. Probably later, the absence of inhibition condition led to solubilisation of more reduce molecules. 3.4. DRIFT spectra of DOM-C and DOM-C-fractions DRIFT spectra of DOM-C, Hi-C and Ho-C for both I-USMSW and S-USMSW are reported in Supplementary data (Fig. S1). The DOM-C spectra for the untreated waste before incubation in simulated landfill (I-USMSW) indicated the presence of a miscellanea of organic compounds. Spectra showed a large band at 3500– 3200 cm1 due to OH– in phenol and carbohydrates and N–H stretching of proteins. The small signal around 2940 cm1 was typical of asymmetric and symmetric stretching of aliphatic C–H. More important was the peaks at 1590 cm1, that was attributed to carboxylate ions, indicating the presence in the DOM-C of volatile fatty acids, and peak at 1430–1424 cm1 due to C–H vibration of aromatic rings (O–CH3). Other peaks were those at 1126 cm1 due to amide I and to C–C–O of ethers. The peaks registered around 1100 cm1 indicated the presence, also, of carbohydrate-like molecules (C–O stretch of carbohydrates; C–C–C stretch of ketone functional group, and –C–OH). Wastes landfilling did not modified DOM-Cs composition as spectra for starting and ending incubation were quite similar. Nevertheless, peaks intensity related to the presence of aliphatic acid, carbohydrates and proteins were less presented than those indicating aromatic-like fraction. Incubation of untreated waste had an effect on DOM-C similar to that of aerobic biostabilization as DRIFT spectra resembled that of DOM-C at the start of incubation (see Fig. S1). On the other hand spectra of S-USMSW before and after incubation were very similar. The spectra of hydrophilic fractions (Hi-C) were quite simple and dominated by peak around 1730 cm1 due to saturated C@O stretch of carboxylic acid, aldehyde, ketone and carboxylic acids. Peak at 1622 cm1 was due to bond deformation of amide I, indicating the presence of protein material; this peaks disappeared with waste incubation. The large peaks at 1200 cm1 was attributed to the stretch of C–O–C in ester. The spectra of Ho-C fractions (Ho-C) were, above all, characterized for the strong peak at 1720 cm1 due to C@O stretching. This peak decreased dramatically with the incubation of both wastes studied. Other peaks were those around of 1200 cm1 due to the stretching of C–O–C of ester and those at 1190–1128 cm1 due to the presence of carbohydrates. As Hi-C and Ho-C contributed to the DOM-C, peaks found in these two fraction were found, also, in the DOM-C spectra (Fig. S1). Nevertheless DOM-C was characterized, also, by peaks that were not described in Hi-C e Ho-C spectra indicating the presence of aromatic moieties, above all in samples after landfill

119

incubation, i.e. peak at 1430–1424 cm1 due to C–H vibration of aromatic rings (O–CH3). NHo-C fraction was responsible for the aromatic moieties characterizing DOM-C in agreement with previous study (e.g. D’Imporzano and Adani, 2007). 3.5. DOM evolution vs. landfill incubation 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 waste getting a high degree of biological stability, i.e. DRI = 978 mg O2 kg SV1 h1 (Salati et al., 2013). Organic matter degradation led to a strong decrease of total DOM-C content of treated waste according to previous work (D’Imporzano and Adani, 2007). Successive incubation of untreated and treated wastes in simulated landfill, determined a strong reduction of DOM-C for both wastes. Nevertheless, if for treated waste DOC decreasing was correlated with TOC reduction (r = 0.96; p < 0.05, n = 4), this was not true for the untreated waste for which no correlation with TOC was found. In particular form Table 4, it could be seen that for untreated waste, TOC at the end of the incubation was equal to that of starting waste indicating that no waste degradation occurred at a large extent. Therefore it can be supposed that DOM-C decreasing for the two incubated wastes was due to different causes. For the S-USMSW, DOM-C decreasing was due to degradation and stabilization of waste under anaerobic condition so that no more hydrolysable fraction was produced and so no new DOM-C occurred (Said-Pullicino et al., 2007). On the other hand for the I-USMSW, DOM-C decreasing was only apparent probably because of inhibition of any biological activity that did not allow the continuing formation of DOM-C replacing degraded and/or leached DOM-C. Therefore, results obtained indicated that biological treatment of waste before landfilling, greatly altered the successive fate of DOM-C. Low pH (for all the incubation period) and the presence of high VFA content (above all during starting period) registered for the untreated (I-USMSW) waste (Table 4) were those typical of a young landfill characterized by an acid phase during witch any biological process is slow down or stopped (Kjeldsen et al., 2002; Tatsi and Zouboulis, 2002). This fact seems to be in contrast with the detail that VFAs decreased during incubation (Table 4). As no degradation processes occurred, i.e. TOC did not decreased (Table 4), it can be assumed that VFAs were, probably, leached. The same fate occurred for DOM-C (Table 2), such as the good correlation found between this parameter and VFAs confirmed (r = 0.98, p < 0.05, n = 4). Therefore, for untreated waste in landfill inhibition of biological activity did not allow both OM hydrolysis forming new DOM-C and its anaerobic fermentation forming new VFAs. On the other hand biostabilized waste (S-USFMSW) showed different trend. High pH and low VFA contents for all the incubation period, indicated an advanced stage of OM decomposition (Kjeldsen et al., 2002). Under this state, inhibitory conditions were avoided and degradation process proceeded fast reducing DOM-C content because of OM degradation/stabilization. DOM-Cs depletion were due, for both wastes, to degradation of hydrophilic and hydrophobic fractions (Table 2). This trend led to the fact that the DOM-Cs at the end of the incubation were constituted mainly of NHo-C fraction that, in effect, increased on a relative basis (Table 2). In particular NHo-Cs contributed for 65% to the DOM-Cs affecting, also, chemical composition of DOM-Cs that assumed, mainly, an aromatic character (D’Imporzano and Adani, 2007) because of the presence of lignin-like material (Nada et al., 1998). This result agreed with previous study reporting that DOM-C from old landfill is composed, above all, by fulvic acid (Christensen et al., 1998) having, mainly, aromatic character (He et al., 2005). On the other hand these results contrasted with those

120

S. Salati et al. / Bioresource Technology 142 (2013) 115–120

Leenheer et al., 2003 which postulated that, above all, terpenoids contributed to DOM-C of landfill leachetes and not lignin like material, although in that study whole wastes were considered, in contrast with this work in which only the organic fraction mechanical sieved, was considered. 4. Conclusion Aerobic treatment of organic fraction of MSW allowed getting good biological stability so that, successive incubation under anaerobic condition in simulated landfill allowed biological process to continue getting a strong reduction of soluble organic matter (DOM-C) that showed mainly, aromatic character. Incubation of untreated waste gave similar trend, but in this case DOM-C decreasing was only apparent because the inhibition of biological process did not allow DOM-C degraded/leached to be replaced by new DOM-C coming from hydrolysis of OM. 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. 2013.05.049. References Adani, F., Genevini, P., Tambone, F., 1995. A new index of organic matter stability. Compost Sci. Util. 3, 25–37. Adani, F., Tambone, F., Gotti, A., 2004. Biostabilization of municipal solid waste. Waste Manag. 24, 8775–8783. Allen, A., 2001. Containment landfills: the myth of sustainability. Eng. Geol. 60, 3– 19. American Public Health Association, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. APHA, Washington, DC. CEN, 2006. Characterization of waste – sampling of waste materials – framework for the preparation and application of a sampling plan. EN, 14899. CEN, 2008. Solid recovered fuels – methods for sampling. PrEN 15442. European Parliament, ropean Parliament. 1991/31/EC of 26 April 1999 on the landfill waste. Official J. Eur. Commun., L182/1–L182/19. Christensen, J.B., Christensen, T.H., 1999. Complexation of Cd, Ni, and Zn by DOC in polluted groundwater: a comparison of approaches using resin exchange, aquifer material sorption, and computer speciation models (WHAM and MINTEQA2). Environ. Sci. Technol. 33, 3857–3863. Christensen, J.B., Jensen, D.L., Gron, C., Zdenek, F., Christensen, T.H., 1998. Characterization of the dissolved organic carbon in landfill leachate-polluted groundwater. Water Res. 32, 125–135. D’Imporzano, G., Adani, F., 2007. The contribution of water soluble and water insoluble organic fractions to oxigen uptake rate during high rate composting. Biodegradation 18, 203–213. De Gioannis, G., Muntoni, A., Cappai, G., Milia, S., 2009. Landfill gas generation after mechanical biological treatment of municipal solid waste. Estimation of gas generation rate constants. Waste Manage. 29, 1026–1034. Dilling, J., Kaiser, K., 2002. Estimation of the hydrophobic fraction of dissolved organic matter in water samples using UV photometry. Water Res. 36, 5037– 5044.

El-Fadel, M., Bou-Zeida, E., Chahineb, W., Alaylic, B., 2002. Temporal variation of leachate quality from pre-sorted and baled municipal solid waste with high organic and moisture content. Waste Manage. 22, 269–282. European Committee for Standardization, 2006. Characterization of waste sampling of waste materials – framework for the preparation and application of a sampling plan. Reference of the method. EN 2006, 14899. Gazzetta Ufficiale n. 281, Repubblica Italiana, 2010. DECRETO 27 settembre 2010. Definizione dei criteri di ammissibilità dei rifiuti in discarica, in sostituzione di quelli contenuti nel decreto del Ministro dell’ambiente e della tutela del territorio 3 agosto 2005, p. 10A14538. He, P.J., Shao, L.M., Qu, X., Li, G.J., Lee, D.J., 2005. Effects of feed solutions on refuse hydrolysis and landfill leachate characteristics. Chemosphere 59, 837–844. Iannotti, D.A., Grebus, M.E., Toth, B.L., Madden, L.V., Hoitink, H.A.J., 1993. Oxygen respirometry to assess stability and maturity of composted municipal solid waste. J. Environ. Qual. 23, 1177–1183. Kjeldsen, P., Barlaz, M.A., Rooker, A.P., Baun, A., Ledin, A., Christensen, T.H., 2002. Present and long-term composition of MSW landfill leachate: a review. Crit. Rev. Environ. Sci. Technol. 32, 297–336. Leenheer, J.A., 1981. Comprehensive approach to preparative isolation and fractionation of dissolved organic carbon from natural waters and wastewater. Environ. Sci. Technol. 15, 578–587. Leenheer, J.A., Nanny, M.A., McIntyre, C., 2003. Terpenoids as major precursors of dissolved organic matter in landfill leachates, surface water, and groundwater. Environ. Sci. Technol. 37, 2323–2331. Lornage, R., Redon, E., Lagier, T., Hébé, I., Carré, J., 2007. Performance of a low cost MBT prior to landfilling: study of the biological treatment of size reduced MSW without mechanical sorting. Waste Manage. 27, 1755–1764. Nada, A.M.A., Fahmy, Y., Abo-Yousef, H., 1998. Kinetic study of delignification of bagasse with butanol-water organosolv pulping process. J. Sci. Ind. Res. 57, 471–476. Pognani, M., Barrena, R., Font, X., Sánchez, A., 2012. Effect of freezing on the conservation of the biological activity of organic solid wastes. Bioresour. Technol. 104, 832–836. Said-Pullicino, D., Erriquens, F.G., Gigliotti, G., 2007. Changes in the chemical characteristics of water-extractable organic matter during composting and their influence on compost stability and maturity. Bioresour. Technol. 98, 1822– 1831. Salati, S., Scaglia, B., Di Gregorio, A., Carrera, A., Adani, F., 2013. The use of the dynamic Respiration Index to predict the potential MSW-leachate impacts after short term mechanical biological treatment. Bioresour. Technol. 128, 351–358. Scaglia, B., Adani, F., 2008. An index to quantifying the aerobic reactivity of municipal solid wastes and derived waste production. Sci. Total Environ. 394, 183–191. Scaglia, B., Adani, F., 2009. Biodegradability of soil water soluble organic carbon extracted from seven different soils. J. Environ. Sci. 21, 1–6. Scaglia, B., Confalonieri, R., D’Imporzano, G., Adani, F., 2010. Estimating biogas production of biologically treated municipal solid waste. Bioresour. Technol. 101, 945–952. Scaglia, B., Paradisi, L., Adani, F., 2011. Intra- and inter-laboratory in real dynamic respiration index (RDRI) method used to evaluate the potential rate of microbial self heating of solid recovered fuel. Bioresour. Technol. 102, 3591–3594. Shao, Z.H., He, P.J., Zhang, D.Q., Shao, L.M., 2009. Characterization of waterextractable organic matter during the biostabilization of municipal solid waste. J. Hazard. Mater. 164, 1191–1197. Tambone, F., Scaglia, B., Scotti, S., Adani, F., 2011. Effect of biodrying process of municipal solid waste properties. Bioresour. Technol. 102, 7443–7450. Tatsi, A.A., Zouboulis, A.I., 2002. A field investigation of the quantity and quality of leachate from a municipal solid waste landfill in a Mediterranean clime (Thessaloniki, Greece). Adv. Environ. Res. 6, 207–219. Zi-gang, L., Chuan-zhou, B., Xiao-lei, J., 2007. Characteristic of Cd sorption in the copper tailings wasteland soil by amended dissolved organic matter from fresh manure and manure compost. Afr. J. Biotechnol. 6, 227–234. Zsolnay, A., 1996. Dissolved humus in soil waters. In: Piccolo, A. (Ed.), Humic Substances in Terrestrial Ecosystems. Elsevier Science, UK, pp. 171–223.