Co-digestion of intermediate landfill leachate and sewage sludge as a method of leachate utilization

Bioresource Technology 102 (2011) 2563–2571

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Co-digestion of intermediate landfill leachate and sewage sludge as a method of leachate utilization A. Montusiewicz ⇑, M. Lebiocka Lublin University of Technology, Faculty of Environmental Engineering, 20-618 Lublin, Nadbystrzycka St. 40 B, Poland

a r t i c l e

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Article history: Received 30 September 2010 Received in revised form 23 November 2010 Accepted 23 November 2010 Available online 30 November 2010 Keywords: Leachate utilization Intermediate landfill leachate Mixed sewage sludge Anaerobic co-digestion Enhanced biogas production

a b s t r a c t This study examines the co-digestion of intermediate landfill leachate and sewage sludge from a municipal wastewater treatment plant. Application of leachate as a co-fermentation component increased the concentrations of soluble organic compounds (expressed as total organic carbon), ammonium nitrogen, and alkalinity in the digester influents. The biogas yield obtained from the co-fermentation of a 20:1 sewage sludge: intermediate leachate mixture was 1.30 m3 per kg of removed volatile solids (VS), while that from a 10:1 mixture was 1.24 m3 per kg of removed VS. These values exceeded the biogas yield for the sludge alone by 13% and 8%, respectively. The leachate addition influenced the proportion of methane to a minor extent. Increased methane yields of 16.9% and 6.2% per kg of removed VS were found for the two sewage sluge:intermediate leachate mixtures, respectively. Ó 2010 Elsevier Ltd. All rights reserved.

1. Introduction Municipal solid waste (MSW) management is a major problem from the environmental, economic, and social points of view. The production of waste grows both in per capita and overall terms (Warah, 2001) as a result of people’s enhanced lifestyles and industrial development. In 1997, global MSW production was 0.49 billion tons (Suocheng et al., 2001), while in 2006 it expanded to about 2.02 billion tons. By 2011, it is expected to have increased by 37.3% (Anonymous, 2007). Taking into account the current concept of sustainable development, controlled sanitary landfills seem to be the most popular way for disposing of solid waste. The reasons for this choice are mainly economic and include cheaper exploitation (compared to incineration and composting) and low capital costs (Renou et al., 2008). According to the Intergovernmental Panel on Climate Change (IPCC) (2006), the fraction of MSW disposed of in landfills reached 67% in Europe, 61% in America, 63% in Asia, 69% in Africa, and 85% in Oceania. A complex series of biological and physicochemical transformations occur as refuse decomposes in landfills (Kjeldsen et al., 2002). Consequently, the biodegradation of the MSW organic fraction together with rainwater percolation through the waste material leads to the generation of a highly contaminated leachate (Hombach et al., 2003). Leachates are characterized by their composition and volumetric flow rates. Leachate flow rate is closely dependent on the ⇑ Corresponding author. Tel.: +48 81 538 4325; fax: +48 81 5381997. E-mail address: [email protected] (A. Montusiewicz). 0960-8524/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2010.11.105

precipitation and evaporation (which differs significantly under various climate conditions), surface run-off, and infiltration of groundwater percolating through the landfill. Moreover, landfill technology (the use of waterproof covers, base liner systems, etc.), waste morphology (particularly water content), as well as compaction of waste layers affects the volumetric flow rate of the leachate (Renou et al., 2008). The chemical composition of MSW landfill leachate depends primarily on the site operations and management, the refuse characteristics, and internal processes (El-Fadel et al., 2002). Major determinants of leachate composition are, usually, the amount and morphology of the deposited waste, the age of the landfill and the corresponding landfill fermentation stage, the depositing technology, and the environmental conditions, e.g. temperature and rainfall (Bilgili et al., 2007). According to Christensen et al. (2001), leachate constituents can be classified into four groups: dissolved organic matter (including CH4, volatile fatty acids and more refractory compounds, e.g. fulvic-like and humic-like compounds), inorganic macro components, heavy metals (e.g. Cd, Cr, Cu, Pb, Ni, and Zn), and xenobiotic organic compounds (including, among others, a variety of aromatic hydrocarbons, phenols, chlorinated aliphatics, pesticides, and plasticizers). A significant amount of biodegradable organic matter enclosed in waste material dumped at young landfills (less than 5 years old) leads to a rapid anaerobic fermentation process, resulting in the enhanced production of volatile fatty acids (VFA), and their subsequent predomination in the leachate (Welander et al., 1997). As a landfill matures, the degree of solid waste stabilization increases and the refractory, non-biodegradable compounds, such as humic

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and fulvic substances, tend to prevail as components of the leachate organic fraction (Chian and DeWalle, 1976; Kulikowska and Klimiuk, 2008). Variation in leachate composition, both over time and from site to site, is particularly visible in terms of the biochemical oxygen demand/chemical oxygen demand (BOD/COD) ratio, which changes from 0.6–0.8 (Chian and DeWalle, 1976; Robinson, 1995; Chen, 1996) to 0.04 (Chian and DeWalle, 1976), and the con1 centration of N–NHþ (Renou 4 that ranges from 0.2 to 13,000 mg L et al., 2008). Although leachate composition can vary widely, a relationship exists between the landfill age and the characteristics of the organic matter. As a consequence, three types of leachate have been defined according to landfill age: recent (from a landfill less than 5 years old with high leachate biodegradability), intermediate (between 5 and 10 years old with medium biodegradability), and old (more than 10 years old with low biodegradability) (Chian and DeWalle, 1976). A knowledge of leachate composition and the variation in its flow rate is critical for choosing a suitable treatment strategy. Generally, leachate treatment requires various, frequently combined, biological and/or physicochemical methods. Conventional treatment includes: (i) leachate transfer, e.g. co-treatment with domestic sewage (Ceçen and Cakiroglu, 2001) and leachate recirculation back through the tip (Bae et al., 1998); (ii) biological aerobic or anaerobic treatment (Klimiuk et al., 2008); and (iii) physicochemical treatment that applies flotation, coagulation/flocculation, chemical precipitation, adsorption, chemical oxidation, ammonia stripping, ion exchange, and electrochemical methods (Renou et al., 2008). An advanced approach incorporates the membrane technology (Renou et al., 2008; Abbas et al., 2009). Biological methods are commonly used for high-strength landfill effluents, although they are not particularly efficient for removing residual and refractory organic compounds. This requires an integrated treatment for an old (less biodegradable) leachate (Abbas et al., 2009). Among many technological approaches to leachate treatment, co-digestion of leachate and sewage sludge seems to be worth considering. Anaerobic treatment of sewage sludge is traditionally involved in municipal wastewater treatment plants (WWTPs) and leads both to sludge volume reduction and to biogas production. However, operational data have indicated possible reserves of the digesters’ capacity, frequently as much as 30%. It would be profitable to use these extra capacities by introducing additional components to conduct the co-digestion process in the existing anaerobic systems. Co-digestion has a lot of advantages. These include dilution of the toxic substances coming from any of the substrates involved, an improved nutrient balance, synergetic effects on microorganisms, a high digestion rate, and possible detoxification based on the co-metabolism process. Moreover, the addition of suitable organic waste ensures an increase in the digester loads, both hydraulic and organic, favoring a more efficient stabilization and enhanced biogas production (Cecchi et al., 1996). In the last decade, co-digestion of sewage sludge and various other components has been studied and reported in the literature. The research concerned the source-sorted organic fraction of municipal solid waste (Hamzawi et al., 1998; Sosnowski et al., 2003; Krupp et al., 2005), manure (Murto et al., 2004), grease trap sludge (Davidsson et al., 2008), and some industrial organic waste (Murto et al., 2004; Bernat et al., 2008; Luste and Luostarinen, 2010). Among other components, using landfill leachate also seems to be a promising solution due to a high concentration of leachate contaminants, particularly soluble organic matter (expressed as soluble COD, total organic carbon (TOC) and volatile fatty acids (VFA)), which indicates that an anaerobic treatment rather than an aerobic one could be applied. However, there are some issues that have to be considered. Mostly they include the landfill age and the corresponding phase of refuse decomposition, commonly

known as major determinants of leachate quality (El-Fadel et al., 2002). Following this, sewage/leachate co-digestion and its effect on biogas generation should be examined. Until now, the subject area has been studied by Hombach et al. (2003) for leachate with a COD concentration of 20,400 mg L1 and leachate doses not exceeding 20% of the raw sludge volume. Other reports have not been found. It is especially worthwhile to investigate the practical aspects of the co-digestion process which involves less-loaded leachate with medium biodegradability, much lower values of COD (than in young unstable leachate), and a BOD/COD ratio decreased to 0.1–0.2. According to Chian and DeWalle (1997), in the middle phase of the landfill life the main leachate organic compounds were refractory long-chain carbohydrates and/or refractory high molecular weight compounds usually not expected to be efficient for organics removal. Thus, investigating the influence of an intermediate leachate addition to the anaerobic digestion of sewage sludge could extend the subject area both with regard to landfill leachate treatment strategies and to the use of another component increasing the effectiveness of the co-digestion process. In the present study, the anaerobic co-digestion of sewage sludge and intermediate landfill leachate was examined to verify whether leachate originating from landfills aged between 5 and 10 years (with medium biodegradability) could be utilized with such a method and whether this could lead to enhanced biogas production. The influence of intermediate leachate additions on the changes in the characteristics of digester influent (consisting of leachate and mixed sewage sludge) was evaluated. The co-digestion process was analyzed based on the changes in bioreactor operational conditions, the VFA effluent profiles, enhanced biogas production, the efficiencies of organics removal and the quality of the digester supernatant. 2. Methods 2.1. Material characteristics Sewage sludge that included two-source residues was obtained from the Puławy municipal wastewater treatment plant (WWTP), Poland, using primary and secondary treatments. Sludge originating from a gravity thickener (i.e. primary thickened sludge) and from a mechanical belt thickener (i.e. waste thickened sludge) was used as the material for our study. The characteristics of mixed sludge from Puławy WWTP are shown in Table 1. Leachate was sourced from the Rokitno, Poland, municipal solid waste landfill. This landfill is of intermediate (medium) age. The leachate composition is also presented in Table 1 and discussed later in Section 3.1. 2.2. Sample preparation procedure The sludge was sampled once a week in the Puławy municipal WWTP and then provided immediately to the Lublin University of Technology (Poland) laboratory. The primary sludge from a gravity thickener and waste sludge from a mechanical belt thickener were transported in separate containers. Under laboratory conditions, the sludge was mixed at a volume ratio of 60:40 (primary waste sludge:waste sludge), then homogenized, manually screened through a 3 mm screen, and partitioned. The sludge samples were frozen at 25 °C in the laboratory freezer and thawed daily for 12 h at 20 °C in the indoor air. Sludge prepared in this manner was considered as mixed sludge, and fed into the reactor. The freezing/thawing conditions were similar to those used in the studies of Jan et al. (2008). However, some modifications were applied. Preliminary experiments showed that the mixed sludge froze

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Table 1 Characteristics of co-digestion components used during experiments. Parameter

Unit

Mixed sludge from Puławy WWTP Average value

COD BOD5 BOD5/COD TOCa VFAa TS VS pH Alkalinity a N–NHþ 4 a P–PO3 4 a b c

mg L1 mg L1 – mg L1 mg L1 g kg1 g kg1 – mg L1 mg L1 mg L1

c

42,519 – – 720 1710 39.0 28.7 6.36 956 150.8 176.2

Upper/lower 95% mean

Leachate from Rokitno landfill c

45,611/39,427 – – 854/586 2188/1232 42.3/35.7 30.1/27.3 6.44/6.28 982/930 210.5/91.1 217.0/135.4

Average value ± standard deviation 1990 ± 32b 290 ± 5b 0.15 ± 0.003b 761 ± 27b 423 ± 12b 8.55 ± 0.08b 2.5 ± 0.03b 7.84 ± 0.08b 8500 ± 187b 1040 ± 32b 4.9 ± 0.3b

Concentrations found in the supernatant. Concentrations found in averaged sample of leachate. The average values and confidence intervals for the whole research period, i.e. 9 months.

completely at 25 °C and at least 12 h at room temperature were required to completely thaw the samples. Leachate was sampled once as an averaged collected sample taken from a leachate storage tank with a capacity of 25 m3 and transported, as soon as possible, to the laboratory. To ensure the same experimental conditions for the co-digestion components, preparation and storage of the leachate samples were similar to those for the sludge. 2.3. Laboratory installation for co-digestion process The laboratory installation consisted of an anaerobic, completely mixed, hermetic reactor equipped with a gas installation, an influent peristaltic pump (feeding the digester), and storage vessels (for influent and effluent). The anaerobic reactor, with a working volume of 0.04 m3, was inserted into the heating jacket at a stable mesophilic temperature. Mixing was carried out using a mechanical stirrer with a rotational speed of 50 min1. Influent was supplied to the upper part of the digester and effluent was wasted through the bottom by gravity. The biogas installation was attached to the headspace of the reactor. The gas system consisted of pipes linked to the pressure equalization unit and the drum gas meter. The gas installation was equipped with gas valves, a dewatering valve, and a gas sampler with a rubber septum, which enabled insertion of a syringe with a pressure lock. The laboratory installation is shown in Fig. 1.

The third run (R3) arrangement was the same as in R2, but this time the volumetric ratio was set at 10:1 (influent consisted of a mixture of 2 L sludge and 200 mL leachate). The HRT reached 18 days and the hydraulic loading rate was 0.055 day1. 2.5. Analytical methods In the mixed sludge, total chemical oxygen demand (COD), volatile fatty acids (VFA), total solids (TS), volatile solids (VS), alkalinity, and pH level were analyzed once a week. The same schedule was used for determining the values of the parameters that characterized the supernatant (sludge liquid phase) before digestion. These parameters included total organic carbon (TOC), ammonium  nitrogen (N–NHþ 4 ), nitrite and nitrate nitrogen (N–NOx ) and ortho3 phosphate phosphorus (P–PO4 ). The supernatant samples were obtained by centrifuging the sludge at 4000 r/min for 30 min. The leachate composition was determined once, after it arrived at the laboratory. The following parameters were analyzed; BOD5,  3 COD, TOC, TS, VS, alkalinity, pH, N–NHþ 4 , N–NOx , P–PO4 . In the digested sludge, the specified parameters were determined three times a week, in accordance with the timetable adopted. Similarly, the supernatant of the digested sludge was examined using the same schedule. Most analyses were carried out in accordance with the procedures listed in the Polish Standards Methods. Some analyses were performed with a FIASTAR 5000 using FOSS analytical methods.

2.4. Operational set-up An inoculum – a collected digest from a mesophilic anaerobic digester operating at 35–37 °C with a volume of 2500 m3 and a hydraulic retention time (HRT) of about 25 days – for the laboratory reactor was taken from the Puławy WWTP. The study was carried out with the reactor operating at a controlled mesophilic temperature of 35 °C and in semi-flow mode (the digester was supplied regularly once a day). Three runs were conducted and each experiment was continued over 90 days (30 days for a complete exchange of reactor contents after a switch in the inflow character and 60 days for the actual experiment – the measurement phase). In the first run (R1) the reactor was fed daily with 2 L of mixed sludge. Hydraulic retention time (HRT) reached 20 days and the hydraulic loading rate was 0.05 day1. The second run (R2) was conducted following the same schedule. However, the reactor was fed using sludge with a leachate added in a volumetric ratio of 20:1 (the influent consisted of a mixture of 2 L sludge and 100 mL leachate). The HRT reached 19 days and the hydraulic loading rate was 0.053 day1.

Fig. 1. Laboratory installation for co-digestion process. (1) anaerobic reactor, (2) mechanical stirrer, (3) heating jacket, (4) influent peristaltic pump, (5) influent storage vessel, (6) effluent storage vessel, (7) drum gas meter, (8) gaseous installation and gas sampler with a rubber septum, (9) dewatering valve, (10) inlet valve, (11) outlet valve.

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Ammonium was determined according to ISO 11732, nitrite and nitrate according to ISO 13395, and ortho-phosphate in accordance with ISO/FDIS 15681-1. The TOC was determined using the Shimadzu TOC-5050A total organic carbon analyzer. Anaerobic digestion efficiency was controlled by the daily evaluation of the biogas yield and its composition (CH4, CO2 and other gases). Moreover, the volatile solids reduction was evaluated according to the Polish Standards Procedure. Biogas production was determined using a Drum-type Gasmeter TG-Series (Ritter, Germany). The composition of the biogas was measured using a Shimadzu GC 14B gas chromatograph coupled with a thermal conductivity detector (TCD) fitted with glass packed columns. The Porapak Q column was used to determine CH4 and CO2 concentrations. The parameters used for the analysis were as follows – injector 40 °C, column oven 40 °C, detector 60 °C, and current bridge 150 mA. The carrier gas was helium with a flux rate of 40 cm3 min1. Peak areas were determined by the computer integration program (CHROMA X). 3. Results and discussion

tent. Taking into account the values of both the organic (TOC, COD and VFA) and inorganic leachate parameters (Table 1 in Section 2.1), Rokitno’s leachate originated, most probably, from a landfill operating at an initial methanogenic phase. This was indicated by a comparison of the values of its constituents with the average values typical for the methanogenic phase found in German landfills and quoted by Jördening and Winter (2005). These were as follows: COD – 2500 g COD m3, BOD5 – 230 g BOD5 m3, TOC – 660 g 3 TOC m3, pH – 7.6, N–NHþ , and total phosphorus – 4 – 740 g m 6.8 g m3. The methanogenic degradation phase of the Rokitno landfill seems to be confirmed also based on leachate compositions reported by Kjeldsen et al. (2002). Values comparable to the data presented by Jördening and Winter (2005) were noticed. Similarly, Kulikowska and Klimiuk (2008) stated that a low COD concentration (<2000 g COD m3), a low BOD5/COD ratio (<0.4), a high pH value (in this study it was 7.84), and concentrations of nutrients comparable to these observed in Rokitno leachate indicated a landfill operating at methanogenic conditions. It should be noted that concentrations of the parameters reported for Rokitno leachate were also common for landfill leachate of medium age (Chen, 1996; Kjeldsen et al., 2002).

3.1. Leachate characteristics According to Jördening and Winter (2005), the constituents of leachates from MSW German landfills from the 1970s and 1980s were defined for the three characteristic degradation phases: first, the acid phase, with a ratio of BOD5/COD P 0.4; second, the intermediate phase with 0.4 > BOD5/COD > 0.2; and lastly, the methanogenic phase, with BOD5/COD 6 0.2. Those phases differed in respect of both BOD5 and COD values as well as other parameters e.g. pH, TOC, ammonium nitrogen concentration, and total phosphorus con-

3.2. Influent characteristics and operational conditions during the experiments The characteristics of the influent that supplied the reactor in runs R1, R2, and R3 are presented in Figs. 2–4 (the left bars, light-grey color). To clarify the digestion results, effluent characteristics and removal efficiency were also compiled on the same figures.

Fig. 2. Influent and effluent characteristics in successive runs as well as removal efficiency: (a) COD (b) TS and (c) VS.

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Fig. 3. Supernatant characteristics in the digester influent and effluent and removal efficiency: (a) TOC and (b) VFA.

 Fig. 4. Nutrient characteristics in the digester influent and effluent as well as release degree: (a) N–NHþ 4 and (b) P–PO4 .

It should be noted that the average values of pH and alkalinity (as CaCO3) in the influents were, respectively, 6.58 and 852 mg L1 for R1, 6.21 and 1310 mg L1 for R2, and 6.50 and 1742 mg L1 for R3. The analysis of the composition of the influent used in the experiments indicated that the average total COD and VS decreased as the leachate dose increased (Fig. 2a and c). A clear tendency appeared with regard to both average alkalinity and TOC (Fig. 3a). Leachate addition caused increases in these parameters, the values achieved for runs R2 and R3 were more than 1.5 times the levels for R1. Also the average VFA concentrations in the influents increased successively in the runs (Fig. 3b). The average TS content (Fig. 2b) and pH levels showed no obvious trends resulting from changes in the composition of the influent. The nutrient concentrations of the influents visibly changed in the experiments (Fig. 4). There was a more than twofold increase in the average N–NHþ 4 concentration in runs R2 and R3 as compared with R1, while the values of P–PO3 exceeded the level in 4 R1 to a minor extent. It should be noticed that the feed conditions varied throughout the experiments carried out over time. This was attributed both to the changes in the sludge parameters and to the addition of leachate. To evaluate the effect of the leachate addition on influent characteristics, the factors reflecting the ‘‘real influence’’ should be identified. Because of the small doses of leachate added to the sewage sludge, an explanation of the increases in influent concentra-

tions required consideration of only those leachate parameters whose values exceeded those for the sewage sludge (Table 1). Substantially, the addition of intermediate landfill leachate as a co-digestion component enhanced the average content of the soluble organic compounds (expressed as TOC), the alkalinity, and the ammonium nitrogen concentration in the digester influent (Figs. 3a and 4a). Contrarily, higher leachate doses caused a decrease of COD in the influent, although the TOC:COD ratio increased concomitantly from a value of 0.014 in R1, to 0.027 in R2, and 0.025 in R3. The increases noted seem to favor the influents enriched by the leachate due to the higher content of dissolved organic matter. The observed increase of average VFA concentrations in the influent could not be attributed to the addition of leachate as its lower VFA content was less than that of the sewage sludge. Rather, it resulted from the higher VFA values characterizing the sewage sludge supplied to the laboratory and used in the R2 and R3 runs. The minor VS:TS ratio in the intermediate leachate (0.29) indicates a much higher content of inorganic compounds as compared to the sewage sludge (VS:TS ratio of 0.765). This is consistent with the research by Lou et al. (2009) regarding the variations of leachate from landfills of different ages. The authors found that inorganic matter increased as the disposal time increased and the values of the TOC:COD ratio in the leachate decreased. Since the leachate composition is closely dependent on the stabilization of the landfill, other parameters will also change significantly as the

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A. Montusiewicz, M. Lebiocka / Bioresource Technology 102 (2011) 2563–2571 Table 2 Operating conditions for anaerobic digestion. Run

Value

Loading rate refer to: COD

VS

TOC

VFA

1.47 1.53/1.41 1.51 1.53/1.49 1.36 1.40/1.30

0.031 0.034/0.028 0.057 0.064/0.050 0.052 0.061/0.043

0.062 0.066/0.058 0.089 0.091/0.087 0.11 0.12/0.10

kg m3 d1 R1 R2 R3

Average Upper/lower 95% mean Average Upper/lower 95% mean Average Upper/lower 95% mean

2.23 2.38/2.08 2.11 2.12/2.10 2.06 2.11/2.01

waste decomposes. The lower values of the dissolved organic matter (TOC, VFA) of the intermediate leachate, as compared to the leachate from a recent landfill, are supposed to favor anaerobic digestion to a minor extent. It should be noted that the hydraulic loading rate and hydraulic retention time (HRT) differed for specified runs. The changes in the influent volume (resulting from the leachate addition to the mixed sludge) led to successive increases in the hydraulic loading rate for the different runs investigated, from 0.05 day1, through 0.053 day1 to 0.055 day1. The HRT values were 20 days for R1, 19 days for R2, and 18 days for R3. The essential values of the loading rates during the experiments are included in Table 2. The differences in the load values were also attributed to the changes in the influents. The COD loading rates decreased as the leachate dose increased. However, the TOC loading rates evaluated for R2 and R3 exceeded by more than 1.5 times the value for run R1, which was in accordance with the higher concentration of solutes present in the leachate. The highest organic loading rate (OLR) relating to VS was estimated during R2; the least value was found in R3. Generally, a higher organic loading rate and a higher concentration of VS leads to more efficient anaerobic digestion, whereas a decrease in such parameters diminishes the efficiency of volatile solids removal (VSR). This was not confirmed in the present studies (Fig. 2c), where the highest VSR of 38.1% was noted in R3 with the least OLR value of 1.36 kg VS m3 day1. During R1, the observed average VSR was 37.2% with the average OLR being 1.47 kg VS m3 day1. A minor biodegradation effect at the level of 31.4% was achieved in R2 despite the highest OLR value of 1.51 kg VS m3 day1. The VSR fluctuations noted could probably be explained as follows. When a leachate dose of 5% was used, the feed composition changed, sludge retention time diminished and the VSR in R2 decreased significantly. However, in R3 an increase in efficiency was observed despite a higher dose (10%) of leachate. The drop in VSR in R2 could be attributed to a shock reaction of the microorganisms to the change in the digester feed and to a possible toxic influence of the leachate on the digester biomass. Moreover, HRT reduction could also have affected the VSR. The subsequent VSR increase in R3 (despite the lowest HRT) seems to indicate that the biomass was adapted to the altered digester environment. The high influent alkalinity probably favors the formation of the specific buffering conditions beneficial for enhanced biogas production despite the shortened sludge retention time.

an increase in the concentration of organic acids in the digest supernatant. An analysis of the concentrations of the intermediate phase products as well as the methane generated may provide insight in the mechanism of the process and testify to the dynamic balance in the system. Variability of the VFA concentrations in the reactor effluent is shown in Fig. 5. The data suggest that methanogenesis in laboratory conditions occurred more effectively than in a full-scale wastewater treatment plant, a fact illustrated by a subsequent decrease of the VFA concentration in run R1. At the end of the run, the VFA concentration was at the level of 208 mg L1, while the VFA concentration in the digest supernatant from Puławy (where the sludge came from) was on average 940 mg L1. Introduction of leachate as a co-digestion component at the rate of 20:1 (sludge to leachate) caused a clear increase in the VFA concentration in the digest supernatant during the first 30 days of the R2 experiment. This means that organic acids generated in the acidogenic phase were not processed. The high VFA concentration and the increasing biogas production indicate an increase in the rate of hydrolysis. The phenomenon observed could result from an extended share of dissolved, fast-hydrolysing substances introduced with leachate. The subsequent days of experiment R2 showed a decrease in the VFA concentration, which after 60 days was maintained at a consistently low level, a fact testifying to a stable course of the process. Thus the adaptation time was about 2 months. Adaptation of the microbial community in the bioreactor was probably aided by an increase in the share of the microorganisms in the leachate from the landfill operating in the methanogenic phase, whose presence is suggested by Laloui-Carpentier et al. (2006). An advantageous effect of methanogenic leachate recirculation on the degradation of recalcitrant waste in the landfill was demonstrated by Vavilin et al. (2006).

3.3. Volatile fatty acids profiles and biogas yields Anaerobic digestion phases (hydrolysis, acidogenesis, acetogenesis and methanogenesis) in semi-flow mode occur simultaneously, each of them at a particular rate. If the system displays an imbalance between the rate of hydrolysis and acidogenesis and that of the generation of acetic acid and its consumption by methanogens for biogas production, then the end effect may be

Fig. 5. Variability of the VFA concentrations in the reactor effluent during experiments.

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An increase in the leachate dose from 5% to 10% in experiment R3 also caused an increase in the VFA concentration, but only during the first 10 days from the moment of the dose increase. During the successive 20 days an unstable course of VFA characteristics was found. In subsequent days the VFA concentration gradually decreased, but only insignificantly. The research results indicate that the dynamic balance among microorganism species after the introduction of leachate co-substrate occurred later than in the case of a double increase of the leachate dose in the influent mixture. The average biogas yields obtained during the investigations are shown in Fig. 6a. The production per kg of VS removed, which was calculated in run R1 and amounted to 1.15 m3 kg1 VS, marginally exceeded the value of 1.12 m3 kg1 VS suggested by Tchobanoglous et al. (2003) as the maximum yield obtainable from the anaerobic digestion of sewage sludge. High yields achieved in the experiments resulted from the use of frozen/thawed sludge as a digester feed during all runs. According to our previous study (Montusiewicz et al., 2010), freezing/thawing disintegration involved in mixed sewage sludge pre-conditioning (with a volumetric ration of primary:waste sludge of 60:40) led to about 5% higher biogas yields per kg of removed VS as compared to raw sludge. Therefore a conversion rate of 1.05, evaluated on the basis of a global mass balance and defined as a ratio of biogas yields achieved from frozen/thawed and raw sludge, respectively, should be used to recalculate the results and compare them with data reported in the literature. After such a recalculation the biogas production per kg of VS removed in R1 was 1.1 m3 kg1 VS. A comparison of biogas yields per kg of removed VS showed that the average values obtained in runs R2 and R3 exceeded by approximately 10% the level achieved in R1. A similar tendency was noticed with regard to the values based on the removed COD with the biogas production levels for R2 and R3 being 1.43 and 1.22 times higher than that of R1. The differences between the average yield values in R2 and R1, as well as in R2 and R3, were statistically significant. Biogas yields calculated per kg of removed TS differed to a minor degree with the highest value being obtained in R1. The enhanced biogas yields were found both in R2 and R3, but in R3 they were achieved despite the lowest OLR and VS (Table 2). This was consistent with a significant increase in the efficiency of TOC and VFA removal during the runs (Fig. 3). The maximum biogas production achieved in R2 could be attributed to the highest organic rates for both VS and TOC; in R3 slightly lower yields were achieved due to the minor loads mentioned. The possible explana-

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tion is that the addition of leachate increases the degree of solubilization in the digester influent as well as improving the buffering conditions (due to the high leachate alkalinity). This seems to be confirmed by the higher alkalinity and TOC concentrations in the influents during runs R2 and R3 (Fig. 3a). An increased content of the solutes regarded as the substrate makes the influent consisting of sludge and leachate more digestible as compared to sludge only, and results in an enhanced biogas yield. In all the experiments, a comparable biogas composition was observed. However, the highest methane share was noted in R2. The average methane content was 56.9% (R1), 58.2% (R2), and 55.7% (R3), with average differences of just 1.3% (R2–R1) and 1.2% (R3–R1). These results indicate that the addition of set doses of intermediate landfill leachate as a co-substrate for the anaerobic digestion of sewage sludge influenced the share of the biogas components to a minor extent. The average methane yields achieved during the studies are presented in Fig. 6b. The value observed during R1 was 0.65 m3 per kg of VS removed, being higher than that of Hombach et al. (2003), who found a level of 0.46 m3 CH4 per kg of VS destroyed in an anaerobic digester fed with mixed sludge and operated for a retention time of 15 days at a temperature of 35 °C. In comparison, our results show a clearly visible difference that could be attributed to using frozen/thawed samples, a longer retention time and a probably higher content of proteins and fats in the sludge. To compare the results achieved with those reported in the literature, the calculations of ‘‘the real methane yield’’ for sludge without pre-conditioning had to be done. The evaluation showed that the recalculated value was 0.65/1.05 = 0.62 m3 CH4 per kg of VS removed, which was still higher than that given by Hombach et al. (2003). Considering the methane yields achieved in R2 and R3, the recalculated values were, respectively, 0.72 m3 CH4 per kg of VS removed for R2 using a low-strength leachate at volumetric dose of 5% and 0.66 m3 CH4 per kg of VS removed for R3 with a volumetric dose of 10%. Such yields were comparable to those noticed by Hombach et al. (2003) in anaerobic digestion systems supplied with sewage sludge enriched with high-strength landfill leachate at volumetric doses of 7% and 12%. The methane yields reported by these researchers increased with the leachate dose and were, respectively, 0.6 and 0.7 m3 CH4 per kg of VS removed. The present results indicate a lower 5% volumetric dose of intermediate (low-strength) leachate to be more favorable for enhanced methane production during co-digestion of sewage sludge and landfill leachate (R2). In this case, with VS contents in bioreactor

Fig. 6. The yields obtained during runs: (a) biogas yields and (b) methane yields.

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influent comparable with R1 (difference less than 3%) as well as OLR (difference less than 2.7%), biogas yield and methane yield were higher by respectively 13% and 16.9%. A decrease in the biogas yield in R3 should rather be linked with an OLR drop; the VS content, compared to R2, was lower by about 14%, causing the OLR value to go down by 10%. However, one cannot exclude the inhibitory effect of toxic substances occurring in leachate, e.g. benzene, toluene or xylene (the authors’ unpublished data). 3.4. Effluent quality The average characteristics of the effluent from the reactor in the runs R1, R2, and R3 are presented in Figs. 2–4 (the right bars, dark-grey color). The average values of the alkalinity and pH were as follows: 3511 mg L1 and 7.80 for R1, 3739 mg L1 and 7.78 for R2, and 3612 mg L1 and 7.66 for R3. Despite the various loading rates caused by influent changes and the progressive decrease in hydraulic retention time from 20 days, through 19 days to 18 days, the stable operation confirmed by a comparable pH level was achieved even in the runs with 5% and 10% leachate additions. Effluent characteristics varied to some extent. However, it was consistent over time probably due to the buffering capabilities of the reactor. Numerically the observed differences reached a few percent and cannot be treated in absolute terms because of the influent changes. Therefore, the proportional analyses for the effluent, such as VS:TS and TOC:COD, should be considered rather than the values of specific parameters. Moreover, a comparison of these ratios with those achieved for influents should be carried out. The results showed that the effluent VS:TS ratio decreased as the leachate dose increased, being 0.655 for R1, 0.646 for R2, and 0.630 for R3. The calculated differences did not exceed 4%, thus indicating quite similar VS shares in the effluent TS despite the variability of the VS:TS influent ratios – 0.765 for R1, 0.707 for R2, and 0.729 for R3. It is worth noting that the highest difference occurred for R2 rather than R1; the maximum difference in the VS:TS ratio was found for the influent (8.2%) and the minimum difference, for the effluent (1.4%). This was consistent with the highest biogas production during R2. The effluent TOC:COD ratios were 0.0125 for R1, 0.011 for R2, and 0.012 for R3. The digested medium revealed similarly low levels of dissolved organic compounds whereas in the influents, higher and significantly different values were found for the sludge enriched by leachate. The TOC:COD values were almost twice the level observed for the sludge only. The minimal ratio of TOC:COD was shown in the digest in R2. Taking into account that the influent ratio in this run had the maximum value, the ‘most effective uptake and use’’ of the soluble organic fraction by biomass could be assumed, thus a highly enhanced biogas production resulted. A similar tendency was observed for R3. However, minor differences between the TOC:COD ratios of the influent and effluent, as well as minor biogas production, were observed as compared to R2. Despite the variability of the nutrients in the influent, quite consistent effluent quality was achieved (Fig. 4). In contrast, a clearly visible release of ammonium nitrogen was found, confirmed by the release degrees (defined as a ratio of the effluent load to influent load) of 5.2 in R1, 2.0 in R2, and 3.1 in R3. Such releases did not appear for the ortho-phosphate phosphorus (the release degree in R1 was 1.03). In the R2 and R3 runs a decrease of P–PO3 4 concentrations during anaerobic digestion occurred and the release degree was only 0.71. The analysis of the nitrogen loads indicates that ammonification released the load of 916 mg d1 in R1, 548 mg d1 in R2 and 983 mg d1 in R3. One of the reasons of this lack of regularity could be the fact that the released ammonium nitrogen could have been precipitated as struvit (MgNH4PO4). Since phosphate concentration in the influent was the greatest in

R2 (Fig. 4b), the amount of the struvit precipitated could have been the maximum one. The observed N–NHþ 4 values in the effluent supernatants were similar to those reported by Song et al. (2004) during mesophilic anaerobic digestion of sewage sludge. Song et al. found that the 1 N–NHþ , whereas during our investi4 concentration was 630 mg L gations the observed average levels were 566.9 mg L1 for R1, 510.8 mg L1 for R2, and 662.2 mg L1 for R3. For the P–PO3 4 concentrations, much lower values were achieved by Song et al. (42.4 mg L1) as compared to our observations of 139.1 mg L1 for R1, 146.6 mg L1 for R2, and 114.5 mg L1 for R3. 4. Conclusions The results indicate that measured additions of intermediate landfill leachate as a co-substrate for the anaerobic digestion of sewage sludge enhanced both biogas and methane yields and influenced the share of biogas components to a minor extent. The best results were found in the experiment using sludge mixed with intermediate leachate in a volumetric ratio of 20:1; however an adaptation time of about two months was needed. The digesters achieved stable operation, so co-fermentation of sewage sludge and intermediate landfill leachate could be considered as a method for such leachate utilization. Acknowledgements The authors acknowledge and are grateful for the financial support of the Ministry of Science and Higher Education, Poland, PBZMEiN-3/2/2006. References Abbas, A.A., Guo, J., Liu, Z.P., Pan, Y.Y., Al-Rekabi, W.S., 2009. Review on landfill leachate treatments. Am. J. Appl. Sci. 6 (4), 672–684. Anonymous, 2007. Global Waste Management Market Assessment, (). Bae, J.H., Cho, K.W., Bum, B.S., Lee, S.J., Yoon, B.H., 1998. Effects of leachate recycle and anaerobic digester sludge recycle on the methane production from solid waste. Water Sci. Technol. 38 (2), 159–168. Bernat, K., Białowiec, A., Wojnowska-Baryła, I., 2008. Co-fermentation of sewage sludge and waste from oil production. Arch. Environ. Prot. 34 (3), 103–114. Bilgili, M.S., Demir, A., Özkaya, B., 2007. Influence of leachate recirculation on aerobic and anaerobic decomposition of solid wastes. J. Hazard. Mater. 143, 177–183. Cecchi, F., Pavan, P., Mata-Alvarez, J., 1996. Anaerobic co-digestion of sewage sludge: application to the macroalgae from the Venice lagoon. Resour. Conserv. Recy. 17, 57–66. Ceçen, F., Cakiroglu, D., 2001. Impact of landfill leachate on the co-treatment of domestic wastewater. Biotechnol. Lett. 23, 821–826. Chen, P.H., 1996. Assessment of leachates from sanitary landfills: impact of age, rainfall and treatment. Environ. Int. 22 (2), 225–237. Chian, E.S.K., DeWalle, F.B., 1976. Sanitary landfill leachates and their treatment. J. Environ. Eng. Div. ASCE 102, 411–431. Chian, E.S.K., DeWalle, F.B., 1997. Evaluation of leachate treatment. Characterization of Leachate, vol. 1. United States Environmental Protection Agency, Cincinnati. Christensen, T.H., Kjeldesen, P., Bjerg, P.L., Jensen, D.L., Christensen, J.B., Baun, A., Albrechtsen, H.J., Heron, G., 2001. Biogeochemistry of landfill leachate plumes. Appl. Geochem. 16, 659–718. Davidsson, Å., Lövstedt, C., la Cour Jansen, J., Gruvberger, C., Aspegren, H., 2008. Codigestion of grease trap sludge and sewage sludge. Waste Manage. 28, 986–992. El-Fadel, M., Bou-Zeid, E., Chahine, W., Alayli, 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. Hamzawi, N., Kennedy, K.J., McLean, D.D., 1998. Anaerobic digestion of co-mingled municipal solid waste and sewage sludge. Water Sci. Technol. 38 (2), 127–132. Hombach, S.T., Oleszkiewicz, J.A., Lagasse, P., Amy, L.B., Zaleski, A.A., Smyrski, K., 2003. Impact of landfill leachate on anaerobic digestion of sewage sludge. Environ. Technol. 24, 553–560. International Panel on Climate Change (IPCC), 2006. Guidelines for National Greenhouse Gas Inventories vol. 5 – Waste (). Jan, T.-W., Adav, S.S., Lee, D.J., Wu, R.M., Su, A., Tay, J.-H., 2008. Hydrogen fermentation and methane production from sludge with pretreatments. Energ. Fuel 22, 98–102.

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