Biodegradation and flushing of MBT wastes

Waste Management 33 (2013) 2257–2266

Contents lists available at ScienceDirect

Waste Management journal homepage: www.elsevier.com/locate/wasman

Biodegradation and flushing of MBT wastes A.A. Siddiqui a,⇑, D.J. Richards b, W. Powrie b a b

Department of Civil Engineering, Aligarh Muslim University, Aligarh 202002, India Waste Management Research Group, Faculty of Engineering and the Environment, University of Southampton, Southampton SO17 1BJ, UK

a r t i c l e

i n f o

Article history: Received 3 October 2012 Accepted 18 July 2013 Available online 20 August 2013 Keywords: MBT waste Biodegradation Flushing Leachate Landfill Pretreatment

a b s t r a c t Mechanical–biological treatment (MBT) processes are increasingly being adopted as a means of diverting biodegradable municipal waste (BMW) from landfill, for example to comply with the EU Landfill Directive. However, there is considerable uncertainty concerning the residual pollution potential of such wastes. This paper presents the results of laboratory experiments on two different MBT waste residues, carried out to investigate the remaining potential for the generation of greenhouse gases and the flushing of contaminants from these materials when landfilled. The potential for gas generation was found to be between 8% and 20% of that for raw MSW. Pretreatment of the waste reduced the potential for the release of organic carbon, ammoniacal nitrogen, and heavy metal contents into the leachate; and reduced the residual carbon remaining in the waste after final degradation from 320 g/kg dry matter for raw MSW to between 183 and 195 g/kg dry matter for the MBT wastes. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Landfill has been the dominant option for municipal solid waste (MSW) disposal in many parts of the world for over a century. However, even the controlled landfilling of MSW may pose risks to humans and the environment. Leachate containing inorganic and organic pollutants may leak into and contaminate the groundwater environment, and degradation gases contribute to global warming if released to the atmosphere. In Europe, the EU Landfill Directive (EC, 1999) has imposed limits on the amount of biodegradable MSW that may be disposed of to landfill. Along with combustion, mechanical biological treatment (MBT) processes are increasingly being adopted as a means of diverting biodegradable municipal waste from landfill in compliance with the Landfill Directive. Mechanical biological treatment prior to landfilling will have major implications for the long term pollution potential of the residue and for future landfill management. The long term behaviour of MBT waste in landfills will be different from that of unprocessed MSW owing to the removal of certain waste fractions during mechanical processing and partial degradation during biological treatment. Knowledge of the gas generation and leaching potential of MBT wastes is needed to assess the risks posed by the receiving landfills, and for their effective design, operation and aftercare including gas and leachate management systems. One of the difficulties with the term MBT is that it covers a wide range of processes, the outputs from which may be very different ⇑ Corresponding author. Tel.: +91 9760249915; fax: +91 5712702783. E-mail address: [email protected] (A.A. Siddiqui). 0956-053X/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.wasman.2013.07.024

even if the input waste is similar. Generally, the mechanical phase of a treatment consists of a combination of sorting, separation, shredding and screening with the purpose of maximising the recovery of recyclables and conditioning the remaining waste for subsequent biological processing. Biological treatment may be anaerobic, aerobic or both, and its aim is to reduce the biodegradability of the residual material which is usually then either burned as an RDF or landfilled. MBT residue from MSW that has not been source-separated is unlikely to be suitable, and may not in any case comply with regulations, for use as a fertilizer or a soil conditioner (Juniper, 2005; BSI/WRAP, 2005). The regulations regarding the landfilling of MBT waste vary across Europe: in Germany, a Landfill Ordinance (German EPA, 2001) sets strict criteria while in the UK there are no equivalent standards. Thus landfilled MBT wastes may vary in their polluting potential as a result of differences in the degree of mechanical sorting and bioprocessing and in the regulations governing their disposal. Robinson et al. (2005) investigated the impact of biological pretreatment on leachate quality and Grilli et al. (2012) demonstrated that leachate recirculation could be effective in encouraging the biostabilization of dried fine fraction MSW after landfilling. However, there is generally relatively little documented experience on the performance of MBT landfills. Previous studies (e.g. Bayard et al., 2008; Bockreis and Steinberg, 2005; van Praagh et al., 2009; Horing et al., 1999; Leikam and Stegmann, 1999) have shown that pretreatment will reduce the gas generation potential and leachate strength of the residue. A different concern is whether the timescale of gas production is similarly affected, and there is conflicting evidence

2258

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

on this. Knox and Robinson (2007) found that gaseous emissions from MBT residue continued over the same time period as for untreated wastes. Conversely, Komilis et al., 1999 and Mahar et al., 2007 found that methanogenic conditions were enhanced by biological pretreatment, leading to a reduction in the time period of gas production. The discrepancy may be a result of differences between field and laboratory conditions relating to water content, mobility and access to the waste. Kylefors et al. (2003) found that landfill simulator reactors gave a better long term prediction of leachate quality than small scale shaking (batch) tests. Leachate constituents of concern in assessing the long term pollution potential of landfills include the total organic carbon (TOC), ammoniacal nitrogen (NH4–N) and heavy metals (Kjeldsen et al., 2002; Price et al., 2003; El-Fadel et al., 2002). This paper presents and discusses the results of a study carried out to investigate the effect of different levels of pretreatment on the gas generating potential and leaching behaviour of MBT wastes. The timescale for gas production and stabilisation of MBT waste is evaluated, and compared with those for untreated MSW in similar conditions. Carbon and nitrogen mass balances are presented, and the extent to which these substances are degraded, leach out of or remain contained within the waste is discussed in the context of the degree of waste treatment.

Table 1 MBT waste composition by material type as dry weight percentages.

2. Materials and methods

2.3. Reactor set-up and operation

2.1. Waste samples

Approximately 40 kg of each waste (obtained from the full sample by coning and quartering) was loaded into a consolidating anaerobic reactor (CAR: Fig. 1) and allowed to degrade anaerobically under a constant vertical load of 50 kPa, simulating anaerobic conditions at a depth of 5 m within a landfill. The body of each CAR comprises a gas-tight Perspex cylinder, 480 mm in diameter and 900 mm heigh. A 100 mm thick gravel drainage layer was placed at the base of the CAR, overlain by a geotextile. MBT waste (which had been oven dried at 70 °C) was then placed into the CAR in layers 50 mm thick to a total height of 500 mm. A further layer of gravel 50 mm thick, was placed over the upper surface of the waste, again separated from the waste by a geotextile. Valves were installed for gas and leachate flow control. The CAR was then sealed and sparged with nitrogen gas to remove any trapped oxygen from the waste, gravel and headspace. 80 l Of laboratory prepared synthetic leachate containing mineral nutrients and trace elements dissolved in deionised water as described by Florencio et al. (1995) was introduced into the waste from the base of the CAR up. To ensure the presence of viable methanogenic bacteria and to accelerate the initiation of methanogenesis, an anaerobically digested sewage sludge from Millbrook Sewage Works, (Southern Water, UK) was added to the synthetic leachate in the volumetric ratioone part sewage sludge to nine parts synthetic leachate (10% by total volume). The CAR and its contents were again sparged with nitrogen gas to remove any remaining oxygen. A vertical stress of 50 kPa was applied to the waste. Each reactor was kept at a steady temperature of approximately 30 °C by means

Two different types of wastes were investigated as follows: (a) Waste MBT-A1: Approximately 500 kg of processed waste was recovered from a mechanical–biological treatment facility in Southern England. The mechanical stage involved removal of bulky recyclable materials, shredding and screening of the remaining waste and ferrous metals recovery. Biodegradation was in forced aerated windrows with regular wetting and turning in an enclosed hall for a period of 6 weeks. The waste was then screened again and the dry recyclables were removed, giving a maximum particle size for the residual waste of about 20 mm. (b) Waste MBT-A2: Approximately 120 kg of processed waste was obtained from an MBT facility situated near Hannover in Northern Germany. The main difference between the two treatment processes was in the biological stage, in which the waste MBT-A2 was anaerobically digested for 3 weeks before being aerobically composted in enclosed windrows for 6 weeks. The maximum particle size of the residual material was about 60 mm. 2.2. Waste composition The composition of each waste by material type (expressed as a percentage of the total dry mass) is given in Table 1. Waste samples were sorted manually by visual identification into material categories of flexible plastics, rigid plastics, textiles, paper, glass, wood, bones, rubber, ceramics, metals and stones. Almost 60% by mass of MBT-A1 and 54% of MBT-A2 could not be identified visually, either because it was too dirty or mixed (>5 mm) or because it was too small (<5 mm). 56% By mass of the sub-63 lm unidentifiable material (and by inference 30% of the total) was lost on ignition, indicating that it was organic in nature. Elemental analysis of the remaining 44% showed the principal components to be SiO2 (23% or 13% of the whole), CaO (10%/6%), Al2O3 (5%/3%) and Fe2O3 (2%/1%). Differences between the two waste compositions were generally insignificant in absolute terms.

Material

Paper Flexible plastics Rigid plastics Wood Textile Rubber Bones Metal Ceramics Stones Glass Unidentified >5 mm Unidentified <5 mm Total

Percent dry weight (%) MBT-A1 waste

MBT-A2 waste

0.43 4.57 6.27 1.57 1.33 0.18 0.27 0.49 2.29 1.73 22.77 28.95 29.15 100.00

0.18 2.4 5.91 3.22 0.63 0.25 0.37 1.49 4.25 3.17 24.36 26.75 27.02 100.00

Waste samples were also analysed for loss on ignition (LOI), total carbon (TC), total nitrogen (TN), cellulose, hemicellulose and lignin contents as discussed in section 2.4; these results are presented in Table 2.

Table 2 Elemental and fibre analysis data for the pretreated wastes. Chemical analysis

Cellulose, % Hemicellulose, % Lignin, % (C + H)/L ratio Total carbon, % Total nitrogen, % Loss on ignition, %

Concentration (% dry mass) MBT-A1 waste

MBT-A2 waste

10.24 4.54 12.63 1.17 22.68 1.81 42.91

7.96 3.91 13.01 0.91 19.85 1.52 34.84

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

2259

2.4. Analytical methods Hydraulic cylinder

Biogas vent

160 mm

Data logger Biogas Leachate pond

Perforated platen

Geotextile 935 mm

Thermocouple

MBT Waste

Leachate recycle

Gravel

100 mm

Geotextile Gravel 480 mm internal diameter

Peristaltic pump

Fig. 1. Schematic view of consolidating anaerobic reactor (CAR).

of a heat blanket, and leachate was recirculated continuously through the reactor using a peristaltic pump. Operation in this way continued for 347 and 279 days for MBT-A1 and MBT-A2 respectively. The volume of biogas produced was determined by allowing the gas to build up in the headspace (of volume Vh) to a small positive pressure (Dp) above ambient atmospheric pressure (pa) as measured by a pressure sensor. The biogas within the reactor was vented automatically to atmosphere through a solenoid valve triggered by the data logging system in response to the limiting pressure rise Dp measured by the pressure sensor. The data logging system recorded each time the valve was activated. The volume of gas (Vg) released at each venting event was then calculated using the ideal gas equations as

Vg ¼

  Dp  Vh pa

ð1Þ

and corrected to dry gas at standard temperature and pressure (STP) using ambient temperature and pressure measurements and values of water vapour pressure as described in Ivanova et al. (2008a). Biogas volume, biogas composition, and leachate quality were monitored regularly as indicators of the progress of waste degradation. Leachate samples were collected from each CAR every three days during the first three months of operation and weekly thereafter. These samples were analysed for pH, electrical conductivity (EC), redox potential (ORP), total organic carbon (TOC), dissolved organic carbon (DOC), inorganic carbon (IC), volatile fatty acids (VFA), total nitrogen (TN), ammoniacal nitrogen (NH4–N), heavy metals, chloride, calcium and magnesium ions. Samples were analysed immediately for pH, EC and ORP, and then preserved and stored at 4 °C until being analysed for the other parameters. The leachate withdrawn from the reactors at each sampling was replaced by an equivalent volume of fresh synthetic leachate. The volume of biogas produced was recorded daily and analysed for gas composition.

The biogas composition (% by total volume of methane and carbon dioxide) was measured daily using a GASCARD II Plus infra-red gas analyser (Edinburgh Instruments, UK). Leachate samples were analysed for pH using a Jenway model 3010 digital pH meter (Jenway, UK). EC and ORP were analysed using portable meters models HI 99301 and HI 8424 (Hanna Instruments, USA) respectively. TOC, DOC, IC and TN analyses of leachate samples were carried out using a high temperature DC-190 TOC analyser (Dohrmann-Rosemount, USA) equipped with a Dohrmann ozonator for TN analysis. NH4–N was measured by steam distillation using a Kjeltec System 1002 distilling unit (Foss Tecator, Sweden). Volatile fatty acids (VFA) were determined by gas chromatography using a Shimadzu-2010 GC (Shimadzu, Japan). Heavy metals, calcium and magnesium were analysed using a Varian Spectra AA-200 atomic absorption spectrophotometer (Varian, Australia). The chloride content in the leachate was measured using Visocolor ECO test kits and a photometer PF-11 (Marcherey-Nagel, Germany). All the leachate samples were analysed in duplicate and the results presented are the average of the two measurements. Representative samples (100 g) of each MBT waste containing the same proportions of each component as the bulk material (Table 1) were prepared and analysed for LOI and TC, TN, cellulose, hemicellulose and lignin contents. Prior to the analyses for LOI, TC and TN, all non grindables (metal, glass, ceramic and stone) were removed from the sample. The remaining waste was dried at 70 °C and milled to a fine powder using a Foss Knifetec 1095 mill in conjunction with a Foss Cyclotec 1093 mill. LOI content was measured by ignition of a dried sample at 550 °C in a muffle furnace (Carbolite, UK) for 2 h. TC and TN contents of the samples were measured using a CE Instruments Flash EA 1112 Elemental Analyser (Thermo Finnigan, Italy). Fibre analysis was carried out using the Foss Analytical FibreCap 2021/2023 system (Kitcherside et al., 2000; Van Soest et al., 1991). Cellulose, hemicellulose and lignin content were determined following the removal of non biodegradable materials (plastic, metal, glass, ceramic and stone), and the drying and milling of the remaining sample as already described. The test comprises three separate procedures of acid detergent fibre (ADF), neutral detergent fibre (NDF) and acid digestible lignin (ADL) analysis to segregate cellulose, hemicellulose and lignin. Each analysis includes procedures of digestion in heated solutions, drying the residue and determining the ash content. These tests were explained and used successfully by Zheng et al. (2007) and Ivanova et al. (2008b) both to characterize the waste samples and assess waste biodegradability. All solids analyses were performed in triplicate and average values are reported.

3. Results and discussion 3.1. Biogas production and leachate pH, VFA and carbon 3.1.1. Biogas, pH and VFA Biogas production and the leachate pH, VFA and carbon contents are all indicators of the occurrence and progress of biodegradation. Data illustrating the evolution over time of the total amount of biogas produced, the leachate pH, total VFA and the total organic carbon (TOC) contents of the leachate are shown in Figs. 2 and 3 for the MBT-A1 and MBT-A2 wastes respectively. Comparative data for raw MSW from the same source tested in the same apparatus by Ivanova et al. (2008a) are given in Fig. 4. Figs. 2 and 3 show that, during the first seven days or so of CAR operation, the total VFA and TOC contents of the leachate peaked and very little biogas was produced. The increase in VFA

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

3000

(a) 40

2000

Biogas TOC

30 20

1000

10 0

Cumulative gas: litre/kg DM

50

20 Total Organic Carbon: mg/litre.

Cumulative gas: litre/kg DM

60

0 0

100

200 Time: days

300

1500

(a) 15

TOC

10

500 5

0

400

0 0

100

200

300

Time: days 9

2000

1000

9

Total VFA

Total VFA

pH

pH

800

8 pH

1000

(b)

8

600 pH

(b)

VFA: mg/litre.

1500

VFA: mg/litre

1000

Biogas

Total Organic Carbon: mg/litre.

2260

400 7

7 500

200

6

0

0

100

200

300

400

0

6 0

100

200

300

Time: days

Time: days Fig. 2. (a) Cumulative biogas production at STP and leachate total organic carbon and (b) leachate VFA and pH for MBT-A1 waste.

concentration was a result of the accumulation of hydrolysis products during a brief initial acidogenic stage of degradation, and was accompanied by a reduction in leachate pH from 7.3 to 6.9. The higher VFA and TOC concentrations in MBT-A1 than in MBT-A2 are consistent with its greater initial organic content. After the first week, biogas production (composition 60% CH4 and 40% CO2 by volume, data not shown) started to increase and the VFA and TOC concentrations in the leachate started to fall, indicating the onset of methanogenesis. The very short acidogenic stage for the MBT wastes in comparison with the 40 days observed for raw MSW by Ivanova et al. (2008a) is consistent with the degradation of organic compounds during MBT (Bayard et al., 2008; Bockreis et al., 2003. However, it is interesting that the absence of deliberate prior anaerobic degradation during the treatment of MBT-A1 did not delay the onset of methanogenesis in the CAR in comparison with MBT-A2, which had been subjected to both aerobic and anaerobic degradation during MBT. The gas production rate increased to a maximum of about 0.9 l/ kg dry matter (DM)/day and 0.4 l/kg DM/day for the MBT-A1 and MBT-A2 wastes respectively, followed by a gradual decline (Figs. 2 and 3). This decrease in biogas production was associated with a gradual reduction in VFA content. Metabolization of the VFA by methanogenic bacteria to methane and carbon dioxide resulted in recovery in the pH. Daily gas production rates had decreased to less than 0.01 litre/kg DM by day 200 for MBT-A1 and by day 150 for MBT-A2. By days 280 and 195 respectively, it had effectively ceased. This was much shorter than the gas production period for raw MSW of about 500 days observed by Ivanova et al. (2008a; Fig. 4). Final stable values of pH were between 7.5 and

Fig. 3. (a) Cumulative biogas production at STP and leachate total organic carbon and (b) leachate VFA and pH for MBT-A2 waste.

7.7, similar to those found in previous studies (e.g. Valencia et al., 2009; Ag˘dag˘ and Sponza, 2005; Warith, 2002). The low final VFA concentrations indicate that most of the available organic matter has been converted into biogas and that biological stabilisation has been achieved . Changes in redox potential and electrical conductivity over time for the leachate in the two MBT wastes are reported by Siddiqui et al. (2012). The redox potential reached about 300 mV indicating strong reducing conditions. The conductivity increased in the first few weeks and then gradually decreased before stabilizing at 13 and 11 mS/cm for MBT-A1 and MBT-A2 respectively (data not shown). The composition of the gas produced by the two wastes was quite similar, with the methane content ranging from 58% to 62% and the carbon dioxide content 35–40%. The total biogas yield was 49.5 l/kg DM for the MBT-A1 and 17.7 l/kg DM for the MBT-A2. The greater gas production from MBT-A1 is consistent with the lesser degree of biological pretreatment, and with the higher organic content indicated by the LOI, cellulose and TC data in Table 2. The biogas yield of MBT-A1 is greater than in some other studies (De Gioannis et al., 2009; Leikam and Stegmann, 1999), but within the range reported by Horing et al. (1999). The volume of biogas produced by raw MSW (255.4 l/kg DM) was significantly greater than either of the two MBT wastes, as was the maximum daily gas generation rate (2–8 l/kg DM) and the gassing time. However, the gas composition was similar. All this is entirely consistent with the greater amount of organic matter remaining in the raw MSW, with a direct correlation between gas potential and LOI or the cellulose + hemicellulose to lignin ratio, (C + H)/L, having been demonstrated by Siddiqui (2011).

2261

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

6000 Biogas TOC

150

4000

100 2000

TOC DOC IC Leachate carbon: mg/litre.

250 200

3000

8000

(a)

Total Organic Carbon: mg/litre.

Cumulative gas: litre/kg DM

300

(a)

2000

1000

50 0 0

200

400 600 Time: days

0

0 1000

800

15000

0

100

200 Time: days

300

3000

8

TOC DOC IC

Total VFA pH 5000

6

0 0

200

400 600 Time: days

800

5 1000

Fig. 4. (a) Cumulative biogas production at STP and leachate total organic carbon and (b) leachate VFA and pH for raw MSW (from Ivanova et al., 2008a).

The reduced gassing potential of the MBT wastes demonstrates diversion of the degradable fraction away from landfill as a result of the biological pretreatment. However, landfill gas control measures will still be needed to prevent fugitive gas emissions. Also, the lower rates and amounts of gas production (8–20% of that for raw MSW) would need to be considered in the design, size and type of gas collection system, and in the economic evaluation of a gas to energy project. There is a possibility that, with only 8–20% of the energy potential of MSW, gas recovery from an MBT landfill would not be economically viable, potentially leading to an increase in fugitive emissions. 3.1.2. Total organic carbon, dissolved organic carbon and inorganic carbon Total organic carbon (TOC) concentrations increased during the first week as a result of the rapid release and hydrolysis of organics from the waste into the leachate during the initial, short acidogenic stage. After the onset of methanogenesis, TOC concentrations began to decrease as biodegradation progressed (mirrored by the increase in gas production and the high methane content of the biogas). TOC concentrations stabilised towards the end of the study, at about 650 and 290 mg/l for the MBT-A1 and MBT-A2 respectively. This residual TOC is mostly recalcitrant (hardly biodegradable) – probably lignin and humic/fulvic acids as suggested by Kjeldsen et al. (2002). The TOC and DOC values were very close (Fig. 5), indicating that most of the organic carbon in the leachate was dissolved. The initial increase in the inorganic carbon (IC) concentration of leachate is probably attributable to the conversion of carbon to inorganic forms such as hydrogencarbonates and carbonates. The subsequent

Leachate carbon: mg/litre.

7

pH

VFA: mg/litre

(b)

10000

400

(b) 2000

1000

0 0

100

200 Time: days

300

400

Fig. 5. Leachate carbon for the (a) MBT-A1 waste (b) MBT-A2 waste.

decrease in IC is due to the precipitation of calcium and magnesium carbonates. The lower organic contents of the leachates associated with the MBT wastes compared with raw MSW is in agreement with the findings of Robinson et al. (2005), Kuehle-Weidemeier (2004) and Horing et al. (1999), but the values are in excess of those observed by Leikam et al. (1999) and van Praagh et al. (2009). The organic strength (i.e., the TOC) of the leachate in MBT-A2 was less than in MBT-A1, consistent with the smaller organic content resulting from the more thorough biological processing during pretreatment (3 weeks of anaerobic and 6 weeks of aerobic biodegradation for MBT-A2 compared with 6 weeks of aerobic biodegradation only for MBT-A1). Fig. 4a shows data of leachate TOC presented by Ivanova et al. (2008a) for raw MSW. Comparison with Figs. 2a and 3a illustrates the effectiveness of pretreatment in reducing the TOC concentration in the leachate, both initially and after stabilization. Thus pretreatment may be beneficial in reducing not only the initial, but also the long term polluting potential of the waste in a landfill. 3.2. Other leachate characteristics: nitrogen, heavy metals, chloride, calcium, and magnesium The significance of these contaminants is that they are relatively unaffected by biodegradation, and must either be removed by flushing or immobilized within the waste. 3.2.1. Ammoniacal and total nitrogen Fig. 6 shows the total (TN) and ammoniacal (NH4–N) nitrogen concentrations in the leachate from the two MBT wastes. Total nitrogen is in principle the sum of nitrite, nitrate, ammoniacal

2262

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

1000

3200 Ammoniacal nitrogen: mg/kg DM.

Total Nitrogen, MBT-A1 Ammoniacal Nitrogen, MBT-A1

Leachate nitrogen: mg/litre

800

Total Nitrogen, MBT-A2 Ammoniacal Nitrogen, MBT-A2

600

400

2800 2400 2000 1600 1200 800 Raw MSW (Ivanova et al. 2008a)

400

MBT-A1

200

MBT-A2

0 0

200

400

600

800

1000

Time: days

0 0

100

200

300

400

Fig. 7. Leachate NH4–N load in raw MSW and pre-treated wastes.

Time: days

Fig. 6. Leachate nitrogen for the MBT-A1 and MBT-A2 wastes.

(Pb), after filtering through Whatman GF/C filters. The changes in concentrations of these heavy metals in the leachates in MBT-A1 and MBT-A2 are shown in Figs. 8 and 9 respectively. After the onset of methanogenesis, metal concentrations tended to decrease owing to the establishment of a highly reducing environment (low redox potential). A significant decrease in all metal concentrations except for chromium and cadmium occurred. Immobilisation of heavy metals from the soluble phase was complete in about 50 days, and concentrations then remained relatively constant until the end of the experiment. Possible metal depletion or immobilisation processes are the formation of insoluble precipitates (principally sulphides or carbonates) and sorption to the waste and suspended solids. Dissolved chromium would be

10

(a)

9

Zn Cu Ni

Concentration: mg/litre.

8 7 6 5 4 3 2 1 0 0

50

100

150

200

250

300

350

400

Time: days

1.0

(b)

0.9

Pb Cr Cd

0.8 Concentration: mg/litre

and organic nitrogen. Since nitrite and nitrate nitrogen levels are expected to be low under anaerobic conditions and a negative redox potential, the total nitrogen in this case comprised mainly ammoniacal and organic nitrogen. Therefore, the total nitrogen (TN) trend was of the same as that for ammoniacal nitrogen. Both TN and NH4–N concentrations increased with time, in contrast to the decrease in TOC. The initial sharp increase in TN and NH4–N may be explained by the direct leaching of ammonia from the waste, and the microbial degradation of nitrogenous organics including proteins and amino acids (Berge et al., 2005; Jokela and Rintala, 2003). After the initial increase, both TN and NH4–N remained stable for a period as reported in studies of landfills with leachate recirculation (Price et al., 2003; Onay and Pohland, 1998). The ammoniacal nitrogen concentration then decreased very slowly to stable values of about 425 mg/l and 195 mg/litre for MBT-A1 and MBTA2 respectively. A similar decrease in ammonia has been observed but not fully explained in bioreactor landfill experiments (Sponza and Ag˘dag˘, 2004; Bilgili et al., 2007). Possible explanations for a decrease in NH4–N are sorption onto the waste mass; anaerobic ammonium oxidation (anammox) and conversion to nitrogen gas (Jun et al., 2009; Zhong et al., 2009; Valencia et al., 2009; Berge et al., 2005); microbial uptake for the growth of new cells; and precipitation of nitrogen as struvite (Kabdasli et al., 2000). The total nitrogen load in the leachate lies within the range previously reported by Horing et al. (1999), but is in excess of that observed by Leikam and Stegmann (1999) and Bayard et al. (2008). The potential for leaching of nitrogen from the MBT-A2 is less than from the MBT-A1, probably as a result of the the lower total nitrogen content of the waste itself. Fig. 7 compares the leachate ammoniacal nitrogen loads measured in the current study on MBT wastes with that obtained by Ivanova et al. (2008a) for raw MSW (data on the leachate total nitrogen content for raw MSW are not available). Not only were the ammoniacal concentrations measured for the leachate from the raw MSW substantially greater, but also there was a marked increase over time in contrast to the gradual decrease observed in the MBT wastes. (The sharp increase in ammoniacal nitrogen concentration just after 600 days was attributed by Ivanova et al. (2008a) to the death/lysis of the bacterial biomass). The benefits of waste pretreatment in reducing leachate ammoniacal nitrogen loads are clear, with the most highly treated waste exhibiting the smallest initial and residual potential pollution loads.

0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0

3.2.2. Heavy metals Leachate samples were analysed for six heavy metals; Zinc (Zn), Nickel (Ni), Copper (Cu), Chromium (Cr), Cadmium (Cd) and Lead

0

50

100

150

200 250 Time: days

300

350

400

Fig. 8. Heavy metals (a) Zn, Cu, Ni and (b) Pb, Cr, Cd in leachate of MBT-A1 waste.

2263

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

4000

10

(a)

9

Zn Cu Ni

3000 Chloride: mg/litre

Concentration: mg/litre.

8 7 6 5 4

2000

1000

3

MBT-A1

2

MBT-A2

1

0 0

0 0

50

100

150

200

250

100

300

200

300

400

Time, days

Time: days

Fig. 10. Chloride in the leachate from the MBT-A1 and MBT-A2 wastes.

1.0

(b)

0.9

Pb Cr Cd

Concentration mg/litre .

0.8

concentration data for MBT-A1 were probably due to interference from bromide ions added as a part of a tracer study. Chloride concentrations are within the range reported by Robinson et al. (2005) and Kjeldsen et al. (2002). The chloride loads (6 g/kg DM for MBT-A1 and 4 g/kg DM for MBT-A2) are at the high end of the range given by Horing et al. (1999), but less than the 10 g/kg DM observed by Boni et al. (2003) for MBT waste.

0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 0

50

100

150

200

250

300

Time: days Fig. 9. Heavy metals (a) Zn, Cu, Ni and (b) Pb, Cr, Cd in leachate of the MBT-A2 waste.

expected to form insoluble hydroxides at the pH and redox potential conditions within the CARs. Heavy metal concentrations were lower in the leachate from MBT-A2 than from MBT-A1. The generally low concentrations of heavy metals are consistent with the results from other studies (van Praagh et al., 2009; Robinson et al., 2005). The average metal concentrations obtained for leachate in raw MSW by Ivanova (2007) were 1.2, 0.45, 0.19, 0.15 and 0.003 mg/litre for Zn, Cu, Ni, Pb and Cd respectively, greater than for the pretreated wastes. Pretreatment of MSW could either have reduced the metal content in the waste, or made the metals less mobileeither of which would be beneficial in terms of reducing pollutant mobility.

3.2.4. Calcium and magnesium The concentrations of calcium (Ca) and magnesium (Mg) in the leachate were determined to facilitate carbon mass balance calculations, presented later. Ca and Mg in the leachate arise from the degradation of organic matter as well as dissolving inorganic waste. Their concentration is controlled by pH and the presence of carbonate ions, which are involved in the precipitation of calcite (CaCO3) and dolomite (Ca Mg (CO3)2) (Christensen et al., 1994). Fig. 11 shows nearly the same trends for calcium and magnesium concentrations in the leachates from both MBT-wastes. The initial high concentrations and the subsequent decrease could indicate that the ions were first leached from the solid refuse and then precipitated as carbonates.

3.3. Carbon and nitrogen mass balance analysis Carbon and nitrogen mass balance calculations – both in terms of the mass of the determinant (carbon or nitrogen) per kg of initial dry matter were carried out in an attempt to track the fate of all of the carbon and nitrogen initially in the reactor.

1500 Calcium, MBT-A1 Calcium, MBT-A2 Magnesium, MBT-A1

Concentration: mg/litre.

3.2.3. Chloride Chloride is one of the major inorganic anions present in the leachate. It is non degradable, conservative and inert, and can only be removed from a landfill by washout. Chloride ion concentration is therefore commonly used to assess leachate dilution (Bilgili et al., 2007), with an increase suggesting that the leachate is becoming more concentrated. Kjeldsen et al. (2002) reported no observable difference in chloride concentration between acidogenic and methanogenic phase leachates owing to the minimal effects of sorption, complexation and precipitation. Fig. 10 shows the chloride concentrations in the leachate from the MBT-A1 and MBTA2 wastes. Chloride ion concentrations increased at the start of the experiment and stayed relatively constant with perhaps a very slow decline. A few slight jumps and fluctuations in the chloride

Magnesium, MBT-A2

1000

500

0 0

100

200

300

400

Time: days Fig. 11. Calcium and magnesium in the leachate from MBT-A1 and MBT-A2 wastes.

2264

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

3.3.1. Carbon mass balance The amounts of carbon in the system at the start and end of each experiment were determined from the measurements of total carbon content (TC of the solid matter measured in mass per kg dry matter directly), and the total organic carbon (TOC) and inorganic carbon (IC) contents of the leachate (measured in mass per unit volume of leachate and multiplied by the leachate volume per kg initial dry matter). During the experiment, carbon exited the waste/leachate system in the vented biogas (CO2 and CH4) and in calcium and magnesium carbonate (CaCO3 and MgCO3) precipitated The carbon lost in the form of methane and carbon dioxide was calculated from the volume of the biogas produced and the volumetric concentrations of CH4 and CO2, normalised per kg of initial dry matter. The loss of carbon due to carbonate precipitates was estimated from the reduction in calcium and magnesium concentrations during the test as explained by Ivanova et al. (2008b). The resulting carbon mass balance can be written as

Cwasteinitially þ Cleachateinitially ¼ CCH4 þ CCO2 þ CCaCO3 þ CMgCO3 þ Cdegradedwaste þ Cleachatefinally

ð2Þ

The mass balance error (expressed as a proportion of the mass of carbon initially in the system) is

waste, 92% of the carbon initially present remained within the waste, about 5% was lost in the biogas (3% methane and 2% carbon dioxide), and less than 0.5% in the leachate. The small amounts of unaccounted for carbon might have deposited as carbonates (calcite, siderite etc.) or biomass in the drainage layer (e.g. Brune et al., 1994; Paksy et al., 1998), or could simply be the cumulative result of small losses during sampling and/or sampling upscaling errors. For raw MSW, Ivanova et al. (2008b) showed that about 70% of the initial carbon remained within the residual waste, 26% was released into the biogas and 1% into leachate after degradation for 919 days. The final carbon contents of the MBT wastes after degradation was complete were 195.1 g/kg initial dry matter for MBT-A1 and 182.9 g/kg initial dry matter for MBT-A2, reduced from immediately post-MBT values of 226.8 g/kg DM and 198.5 g/kg DM respectively. These may be compared with values for raw MSW of 460.5 g/kg DM initially and 318.9 g/kg DM after full degradation (Ivanova et al., 2008b). This would suggest that the pretreatment of MSW will result in the removal of more carbon than leaving it to degrade in the relatively uncontrolled environment of a landfill, and a residual carbon load of 57–63% of that for raw MSW.

  Cwasteinitially þ Cleachate initially  CCH4  CCO2  CCaCO3  CMgCO3  Cdegradedwaste  Cleachate finally Cwasteinitially þ Cleachateinitially The carbon mass balance calculations are summarised in Table 3. The error was about 2% for MBT-A1 and 3% for MBT-A2, which in view of the inherent uncertainty in upscaling the results of small samples to the whole gives a substantial degree of confidence in the data. 11% was removed in the biogas (about 7% as methane and 4% as carbon dioxide), and less than 1% was transferred into the leachate. The mass balance shows that 86% of the carbon initially in MBT-A1 remained there, 11% was removed in the biogas (about 7% as methane and 4% as carbon dioxide), and less than 1% was transferred into the leachate. For the MBT-A2

ð3Þ

3.3.2. Nitrogen mass balance As far as nitrogen is concerned, the system is essentially closed and the mass balance may be written (again, in terms of mass of nitrogen per kg of initial dry matter) as

Nwasteinitially þ Nleachateinitially ¼ Ndegradedwaste þ Nleachatefinally

ð4Þ

and the mass balance error (again expressed as a proportion of the nitrogen initially in the system) is

Table 3 Summary of carbon mass balance.

Carbon initially in system

Carbon removed from the system Carbon remaining in the system

Location

Determined from

MBT-A1 waste g/kg DM

MBT-A2 waste g/kg DM

Carbon in the initial waste, g/kgDM Carbon in synthetic leachate, g/kgDM Carbon in methane, g/kgDM

Measurement of total carbon (TC) in the initial waste

226.8

198.5

1.0

0.8

16.2

5.4

9.5

3.6

2.5

1.6

195.1

182.9

0.6

0.3

0.04

0.1

1.7

2.7

Carbon in carbon dioxide, g/ kgDM Carbon in leachate, g/kgDM

Carbon in the degraded waste, g/kgDM Carbon precipitated as calcium carbonates, g/ kgDM Carbon precipitated as magnesium carbonates, g/ kgDM Mass balance error, %

(Measured TOC + measured IC)  total volume of leachate in CAR  initial dry mass of waste in the CAR Total volume of biogas produced  volumetric proportion of CH4  proportion of total volume attributable to carbon  initial dry mass of waste in the CAR Total volume of biogas produced  volumetric proportion of CO2  proportion of total volume attributable to carbon  initial dry mass of waste in the CAR (Measured TOC + measured IC)  total volume of leachate in CAR  initial dry mass of waste in the CAR Measurement of total carbon (TC)  final dry mass of waste  initial dry mass of waste in the CAR Change in concentration of calcium in leachate  total volume of leachate in CAR  proportion of total mass attributable to carbon  initial dry mass of waste in the CAR Change in concentration of magnesium in leachate  total volume of leachate in CAR  proportion of total mass attributable to carbon  initial dry mass of waste in the CAR

2265

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266 Table 4 Summary of nitrogen mass balance.

Nitrogen initially in system

Nitrogen remaining in the system

Location

Determined from

MBT-A1 waste g/kg DM

MBT-A2 waste g/kg DM

Nitrogen in the initial waste, g/kgDM Nitrogen in synthetic leachate, g/kgDM Nitrogen in the degraded waste, g/kgDM Nitrogen in leachate, g/kgDM Mass balance error, %

Measurement of total nitrogen (TN) in the initial waste

18.1

15.2

0.2

0.1

15.9

14.1

Measured TN  total volume of leachate in CAR  initial dry mass of waste in the CAR Measurement of total nitrogen (TN)  final dry mass of waste  initial dry mass of waste in the CAR Measured TN  total volume of leachate in CAR  initial dry mass of waste in the CAR

  Nwasteinitially þ Nleachate initially  Ndegradedwaste  Nleachate finally Nwasteinitially þ Nleachateinitially

1.24

0.5

6.3

4.9

ð5Þ

Nitrogen mass balance calculations for the MBT-A1 and MBT-A2 wastes are summarised in Table 4, and agree to within about 6% for both. In the case of MBT-A1, 88% of the nitrogen remained in the waste at the end of degradation and 6% transferred into the leachate, with 6% unaccounted for. For the MBT-A2 waste, 93% of the nitrogen remained in the waste, 2% transferred into the leachate and 5% was unaccounted for. The unaccounted nitrogen was probably lost with the gas (of which about 5% was neither methane nor carbon dioxide) and possibly by microbial uptake or struvite precipitation within the drainage layer. Full details of the carbon and nitrogen mass balances are given in Siddiqui (2011).

4. Conclusions 1. Measurements of gas generation and leachate quality in terms of TOC, NH4–N and VFA have demonstrated that in laboratory scale reactors, waste stabilisation was achieved in less than a year for MBT wastes subjected to two quite different degrees of pretreatment. In both cases, the initial acidogenic stage of degradation in simulated landfill conditions was short (less than a week) – even when the pretreatment had been entirely aerobic (MBT-A1). In comparison, the degradation time for raw MSW in similar conditions was more than 400 days (including an initial acidogenic period of about 40 days). 2. The prior treatment of MSW substantially reduces its gas generating potential when the waste is subsequently placed within a landfill. Six weeks of aerobic pretreatment (MBT-A1) led to a 80% reduction in the gas generating potential compared with raw MSW. Adding a 3-week anaerobic stage (MBT-A2) reduced the gas generating potential by about 92%. These low rates of gas production may be problematic for the effective and economic design and operation of gas collection and management systems, and could affect the potential economic viability of energy recovery systems. This is a major potential disbenefit of the partial treatment of biodegradable MSW prior to landfilling. 3. Waste pretreatment resulted in reduced TOC, ammoniacal nitrogen and heavy metals contents in the leachate. This is potentially beneficial in reducing the potential for pollution when the waste is subsequently landfilled, andthe timescale over which active leachate management is required. 4. The carbon and nitrogen mass balances indicated that only a small proportion of each was unaccounted for. A large proportion of both carbon and nitrogen remained in the waste material and was not released. The amount of carbon remaining in the MBT waste after full degradation was 57–63% of that for raw MSW.

References Ag˘dag˘, O.N., Sponza, D.T., 2005. Co-digestion of municipal sludge with municipal solid wastes in anaerobic simulated landfilling reactors. Process Biochemistry 40, 1871–1879. BSI/WRAP, 2005. PAS100:2005 Specification for Composted Materials. British Standards Institution. Bayard, R., de Araujo Morais, J., Rouez Fifi, M.U., Achour, F., Ducom, G., 2008. Effect of biological pretreatment of coarse MSW on landfill behaviour: laboratory study. Water Science and Technology 58 (7), 1361–1369. Berge, N.D., Reinhart, D.R., Townsend, T.G., 2005. The fate of nitrogen in bioreactor landfills. Critical Reviews in Environmental Science and Technology 35 (4), 365– 399. Bilgili, M.S., Demir, A., Öskaya, B., 2007. Influence of leachate recirculation on aerobic and anaerobic decomposition of solid wastes. Journal of Hazardous Materials 143, 177–183. Bockreis, A., Steinberg, I., Rohde, C., Jager, J., 2003. Gaseous emissions of mechanically-biologically pre-treated waste for long term experiments. In: Proceedings Sardinia 2003, Ninth International Waste Management & Landfill Symposium. S. Margherita di Pula, Cagliary, Italy, 6–10 October 2003. Bockreis, A., Steinberg, I., 2005. Infuence of mechanical–biological waste pretreatment methods on the gas formation in landfills. Waste Management 25, 337–343. Boni, M.R., Chiavola, A., Napoleoni, Q., Sbaffoni, S., Stacul, E., 2003. Biochemical and mechanical behaviour of differenet pretreated waste. In: Proceedings Sardinia 2003, Ninth International Waste Management & Landfill Symposium, S. Margherita di Pula, Cagliary, Italy, 6–10 October. Brune, M., Ramke, H.G., Collins, H.J., Hanert, H.H., 1994. Incrustation problems in landfill drainage systems. Landfilling of waste: Barriers, E & FN Spon, London, pp. 569–605. Christensen, T.H., Kjeldsen, P., Albrechtsen, Hans-Jorgen, Heron, G., Neilsen, P.H., Bjerg, P.L., Holm, P.E., 1994. Attenuation of Landfill Leachate Pollutants in Aquifers. Critical Reviews in Environmental Science and Technology 24 (2), 119–202. 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 constants. Waste Management 29, 1026–1034. EC, 1999. Directive 1999/31/EC on the Landfill of Waste. Council of the European Union, Official Journal of the European Communities, vol. L182, pp. 1–19. 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 Management and Research 22, 269. Florencio, L., Field, J.A., Lettinga, G., 1995. Substrate competition between methanogens and acetogens during the degradation of methanol in UASB reactors. Water Research 29 (3), 915–922. German EPA, 2001. Ordinance on Environmentally Compatible Storage of Waste from Human Settlements and on Biological Waste Treatment Facilities (AbfAblV). Federal Ministry for the Environment, Nature Conservation and Nuclear Safety, Berlin,Germany, 2007,
2266

A.A. Siddiqui et al. / Waste Management 33 (2013) 2257–2266

Jokela, J.P.Y., Rintala, J.A., 2003. Anaerobic solubilisation of nitrogen from municipal solid waste. Reviews in Environmental Science & Biotechnology 2, 67–77. Jun, D., Yong-sheng, Z., Hong, Mei, Zhang, W., 2009. Influence of alkalinity on the stabilization of municipal solid waste in anaerobic simulated bioreactor. Journal of Hazardous Materials 163 (2–3), 717–722. Juniper, 2005. Mechanical–Biological Treatment: A Guide for Decision Makers, Processes, Policies and Markets – Annexe A, Process Fundamentals. Juniper Consultancy Services Ltd., Sheppards Mill, UK. Kabdasli, I., Tünay, O., Öztürk, I., Yilmaz, S., 2000. Ammonia removal from young landfill leachate by magnesium ammonium phosphate precipitation and air stripping. Water Science and Technology 41 (1), 237–240. Kitcherside, M.A., Glen, E.F., Webster, A.J.F., 2000. FibreCap: an improved method for the rapid analysis of fibre in feeding stuffs. Animal Feed Science and Technology 86, 125–132. Kjeldsen, P., Barlaz, M.A., Rooker, R., Baun, A., Ledin, A., Christensen, T.H., 2002. Present and long-term composition of MSW landfill leachate: a review. Critical Reviews in Environmental Science and Technology 32 (4), 297–336. Knox, K., Robinson, H.D., 2007. MBT and thermal treatment of MSW residues: a comparative study of energy balance and long-term pollution potential of leachates. In: Proceedings Sardinia 2007, Eleventh International Waste Management and Landfill Symposium S. Margherita di Pula, Cagliari, Italy; 1– 5 October. Komilis, D.P., Ham, R.K., Stegmann, R., 1999. The effect of municipal solid waste pretreatment on landfill behaviour: a literature review. Waste Management and Research 17, 10–19. Kuehle-Weidemeier, M., 2004. Landfill properties of mechanically and biologically treated municipal solid waste. WASTE 2004 Conference Integrated Waste Management and Pollution Control: Policy and Practice, Research and Solutions. Stratford-upon-Avon, Warwickshire, UK. Kylefors, K., Andreas, L., Lagerkvist, A., 2003. A comparison of small-scale, pilotscale and large-scale tests for predicting leaching behavior of landfilled wastes. Waste Management 23, 45–59. Leikam, K., Stegmann, R., 1999. Influence of mechanical–biological treatment of municipal solid waste on landfill behaviour. Waste Management and Research 17 (1999), 424–429. Leikam, K., Jahnke, S., Raga, R., Stegmann, R., 1999. Influence of water content and emplacement density on landfill emissions. In: Christensen, T.H., Cossu, R., Stegmann, R. (Eds.), Proceedings Sardinia 1999, Seventh International Waste

Management and Landfill Symposium, S. Margherita di Pula, Cagliari, Italy, CISA, Environmental Sanitary Engineering Centre, Italy, pp. 495–502. Mahar, R.B., Liu, J., Yue, D., Nie, Y., 2007. Landfilling of pretreated municipal solid waste by natural convection of air and its effects. Journal of Environmental Science and Health Part A 42, 351–359. Onay, T.T., Pohland, F.G., 1998. In situ nitrogen management in controlled bioreactor landfills. Water Research 32, 1383–1392. Paksy, A., Powrie, W., Robinson, J.P., Peeling, L., 1998. A laboratory investigation of anaerobic microbial clogging in granular landfill drainage media. Géotechnique 48 (3), 389–401. Price, G.A., Barlaz, M.A., Hater, G.R., 2003. Nitrogen management in bioreactor landfills. Waste Management 23, 675–688. Robinson, H.D., Knox, K., Bone, B.D., Picken, A., 2005. Leachate quality from landfilled MBT waste. Waste Management 25, 383–391. Siddiqui, A.A., Richards, D.J., Powrie, W., 2012. Investigations into the landfill behaviour of pretreated wastes. Waste Management 32, 1420–1426. Siddiqui, A.A., 2011. Biodegradation and Settlement Characteristics of MBT Waste, PhD Thesis, University of Southampton. Sponza, D.T., Ag˘dag˘, O.N., 2004. Impact of leachate recirculation and recirculation volume on stabilization of municipal solid wastes in simulated anaerobic bioreactors. Process Biochemistry 39, 2157–2165. Valencia, R., van der Zon, W., Woelders, H., Lubberding, H.J., Gijzen, H.J., 2009. Achieving ‘‘Final Storage Quality’’ of municipal solid waste in pilot scale bioreactor landfills. Waste Management 29 (2009), 78–85. van Praagh, M., Heerenklage, J., Smidt, E., Modin, H., Stegmann, R., Persson, K.M., 2009. Potential emissions from two mechanically–biologically pretreated (MBT) wastes. Waste Management 29 (2009), 859–868. Van Soest, P.J., Robertson, J.B., Lewis, B.A., 1991. Methods for dietary fiber, neutraldetergent fiber and nonstarch polysaccharides in relation to animal nutrition. Journal of Dairy Science 74, 3583–3597. Warith, M., 2002. Bioreactor landfills: experimental and field results. Waste Management 22, 7–17. Zheng, B., Richards, D.J., Smallman, D.J., Beaven, R.P., 2007. Assessing MSW degradation by BMP and fibre analysis. Waste and Resource Management 160 (WR4), 133–139. Zhong, Q., Li, D., Tao, Y., Wang, X., He, X., Zhang, J., Guo, W., Wang, L., 2009. Nitrogen removal from landfill leachate via ex situ nitrification and sequential in situ denitrification. Waste Management 29 (4), 1347–1353.