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UNIVERSITY OF WISCONSIN SYSTEM SOLID WASTE RESEARCH PROGRAM

Microbial Diversity and Dynamics during Methane Production from Municipal Solid Waste

2010

Christopher A. Bareither Georgia L. Wolfe Katherine D. McMahon Craig H. Benson

University of Wisconsin-Madison

Waste Management 33 (2013) 1982–1992

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Waste Management

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Microbial diversity and dynamics during methane production from municipal solid waste ⇑ Christopher A. Bareither a,b, , Georgia L. Wolfe c, Katherine D. McMahon d, Craig H. Benson e a Civil & Environmental Engineering, Colorado State University, Ft. Collins, CO 80532, USA b Geological Engineering, University of Wisconsin-Madison, Madison, WI 53706, USA c Bacteriology, University of Wisconsin-Madison, Madison, WI 53706, USA d Bacteriology, Civil and Environmental Engineering, University of Wisconsin-Madison, Madison, WI 53706, USA e Civil & Environmental Engineering, Geological Engineering, University of Wisconsin-Madison, Madison, WI 53706, USA article info abstract

Article history: The objectives of this study were to characterize development of bacterial and archaeal populations dur- Received 1 October 2012 ing biodegradation of municipal solid waste (MSW) and to link specific methanogens to methane gener- Accepted 14 December 2012 ation. Experiments were conducted in three 0.61-m-diameter by 0.90-m-tall laboratory reactors to Available online 12 January 2013 simulate MSW bioreactor landfills. Pyrosequencing of 16S rRNA genes was used to characterize microbial communities in both leachate and solid waste. Microbial assemblages in effluent leachate were similar Keywords: between reactors during peak methane generation. Specific groups within the Bacteroidetes and Ther- Bioreactor landfill matogae phyla were present in all samples and were particularly abundant during peak methane gener- Methane ation. Microbial communities were not similar in leachate and solid fractions assayed at the end of Methanogen Municipal solid waste reactor operation; solid waste contained a more abundant bacterial community of cellulose-degrading Quantitative PCR organisms (e.g., Firmicutes). Specific methanogen populations were assessed using quantitative polymerase chain reaction. Methanomicrobiales, Methanosarcinaceae, and Methanobacteriales were the predominant methanogens in all reactors, with Methanomicrobiales consistently the most abundant. Methanogen growth phases coincided with accelerated methane production, and cumulative methane yield increased with increasing total methanogen abundance. The difference in methanogen populations and corresponding methane yield is attributed to different initial cellulose and hemicellulose contents of the MSW. Higher initial cellulose and hemicellulose contents supported growth of larger methanogen populations that resulted in higher methane yield. Ó 2012 Elsevier Ltd. All rights reserved.

1. Introduction Methane production from MSW depends on a consortium of microorganisms that degrade complex organic molecules in Municipal solid waste (MSW) landfills remain the predominant sequential phases referred to as hydrolysis, fermentation, aceto- end-state for waste disposal in the US (USEPA, 2010). Operation of genesis, and methanogenesis (Farquhar and Rovers, 1973; Zehnder, a MSW landfill as a bioreactor is an attractive waste management 1978; Barlaz et al., 1989; Zinder, 1993; Pohland and Kim, 1999; option for in situ treatment and stabilization of solid waste and Levén et al., 2007). The initial step is hydrolysis of complex poly- leachate (Pohland, 1980; Barlaz et al., 1989; Reinhart et al., 2002; mers (e.g., cellulose, hemicellulose, starch, protein) to lower molec- Mehta et al., 2002; Benson et al., 2007; Bareither et al., 2010a; Bar- ular weight monomers (e.g., sugars, amino acids). Fermentation of laz et al., 2010). Anaerobic decomposition of MSW in landfills these monomers to alcohols, carboxylic acids (e.g., acetate, propio- yields methane (CH4) and carbon dioxide (CO2) as end-products. nate, butyrate), hydrogen (H2), and carbon dioxide (CO2), combined Methane can be collected and used in on-site biogas-to-energy sys- with acetogenesis of H2 and CO2 to acetate, yields the substrates tems, which adds to the US alternative energy portfolio and de- and chemical and microbial equilibrium necessary for methano- creases methane released to the atmosphere as a greenhouse gas. genesis (Zehnder, 1978). Methane producing Archaea (i.e., methanogens) are critical to efficient and balanced waste decomposition as they carry out the terminal step in the anaerobic ⇑ Corresponding author at: Civil & Environmental Engineering, Colorado State decomposition process. University, Ft. Collins, CO 80532, USA. Tel.: +1 9704914660. Methanogenesis occurs primarily through two pathways: E-mail addresses: [email protected] (C.A. Bareither), gwolfe@ wisc.edu (G.L. Wolfe), [email protected] (K.D. McMahon), chbenson@wisc. acetotrophic and hydrogenotrophic decomposition (Zinder, edu (C.H. Benson). 1993). Acetotrophic methanogenesis involves consumption of

0956-053X/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.wasman.2012.12.013 C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992 1983 acetate, whereas hydrogenotrophic methanogenesis involves con- tall stainless steel tanks. The initial waste specimens contained sumption of H2 and CO2 or formate (Zehnder, 1978; Pavlostathis 115 kg of dry MSW and were compacted to initial waste volumes and Giraldo-Gomez, 1991). Hydrogenotrophic methanogens form of approximately 188 L (0.188 m3). Gravel was placed at the base syntrophic associations with hydrogen-producing fermentative of the tank for leachate collection and on top of the waste speci- and acetogenic bacteria, allowing the syntrophic partner to carry men to distribute recirculated leachate. A fine-mesh aluminum out fermentations by removing inhibitory waste products. screen was placed between the waste and drainage gravel to pre- Methanogens pertaining to both acetotrophic and hydrogeno- vent clogging of the effluent port. Leachate was collected in a trophic decomposition pathways have been identified in MSW sealed, inert plastic bag to minimize exposure to oxygen, and landfills (e.g., Huang et al., 2002; Huang et al., 2003; Calli and Cir- was recirculated via perforated PVC pipes installed in the distribu- gin, 2005; Calli et al., 2006; Laloui-Carpentier et al., 2006). Staley tion gravel. Biogas generated during waste decomposition passed et al. (2011a) report that acid-tolerant acetoclastic methanogens through an outlet port in the reactor lid and was collected in 40- may play a crucial role in initiating methanogenesis in MSW and L flexfoil gasbags (SKC Inc., Eight Four, PA). neutralizing pH. Methanogen population dynamics have been Waste temperature was maintained between approximately studied extensively in other anaerobic systems (e.g., wastewater 30–40 °C using two 300-mm 600-mm flexible silicone rubber treatment, anaerobic digesters), yet the environmental, opera- heaters affixed to the sides of the reactors (Fig. 1). Type-T thermo- tional, and material factors promoting dominance of acetotrophic couples (Omega Engineering, Inc., Stamford, CT) were used to mon- vs. hydrogenotrophic methanogenesis are complex and still not itor waste temperatures in the drainage gravel, distribution gravel, well understood (Demirel and Scherer, 2008; Semrau, 2011). and at four vertically staggered locations in the MSW (Fig. 1). Tem- The objectives of this study were to characterize development peratures were recorded and controlled with a Campbell Scientific of microbial communities and quantify methanogen populations CR23X Micrologger (Campbell Scientific Inc., Logan, Utah) con- during MSW decomposition in laboratory-scale bioreactor land- nected with an AM25T thermocouple multiplexer. fills. Microbial communities and methanogen population abundances were linked to temporal changes in leachate chemistry 2.2. Municipal solid waste and methane generation. Three 188-L laboratory reactors (A, B, and C) containing shredded MSW were operated for 181 d at mes- Municipal solid waste was collected from Deer Track Park Land- ophilic temperatures with leachate recirculation. Leachate chemis- fill (Watertown, WI) during construction of a field-scale lysimeter try and methanogen populations were evaluated weekly. experiment to assess bioreactor performance (Bareither et al., Decomposed solid waste was exhumed from one reactor to com- 2012). MSW had been in-place approximately 3–4 months at the pare microbial communities in the solid waste and leachate frac- time of sampling and contained the following composition on a tions. Quantitative polymerase chain reaction (qPCR) targeting dry mass basis: 19.0% paper/cardboard, 17.0% gravel/ceramics/in- 16S rRNA genes was used to assess abundance of specific methano- erts, 7.8% wood, 5.2% flexible plastic, 5.6% metal, 5.6% rigid plastic, gens in leachate and solid waste. Pyrosequencing of 16S rRNA 2.9% textile, 2.8% miscellaneous, 2.9% glass, 1.1% food waste, 0.3% genes was used to describe bacterial community composition. yard waste, and 29.8% soil-like material (i.e., material passing a 4.75-mm sieve). The initial waste composition was the same in all three reactors. 2. Materials and methods MSW constituents with particle diameters greater than 25 mm were shredded and passed through a 25-mm screen prior to place- 2.1. Laboratory reactors ment in the laboratory reactors. Shredded MSW was blended thor- oughly by hand to achieve homogenized waste specimens for all A schematic of a laboratory reactor used in this study is shown three reactors. Waste specimens were hydrated to a dry weight in Fig. 1. The reactors consisted of sealed 0.6-m-diameter by 0.9-m- water content of 28% and compacted in five lifts to an average total unit weight of 7.7 kN/m3. Solid waste chemical characteristics, including cellulose (C), hemicellulose (H), and lignin (L) contents; volatile solids (VS); and biochemical methane potential (BMP); were measured on the initial MSW used to fill the laboratory reactors as well as on MSW exhumed from the reactors following experimentation. A summary of these chemical characteristics is in Table 1. A description of the methods used to determine chemical characteristics of the wastes is in Bareither et al. (2010b). The C and H contents, [(C + H)/L], and BMP all decreased for the exhumed wastes relative to the initial waste, indicating that MSW in the reactors decomposed relative to the initial MSW. Further discussion on the implications of waste composition variability on reactor performance and microbial community development is presented subsequently.

2.3. Leachate management

Leachate used for dosing the reactors was collected from Deer Track Park Landfill (Watertown, WI). A 10-L dose was added on Day 41 and on Day 43 of reactor operation to inoculate the waste with an active anaerobic microbial community. Subsequent leachate doses were executed in 1-L volumes applied every 1–2 d throughout the duration of the experiment. Effluent leachate was Fig. 1. Schematic of a laboratory reactor. Initial waste specimen volume was approximately 188 L (0.188 m3). Circles indicate locations of solid waste exhumed consistently recirculated in the reactors, with additional fresh from Reactor A for 454 pyrosequencing. leachate added if needed to achieve 1 L. A sufficient volume of 1984 C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992

Table 1 Chemical characteristics of the initial municipal solid waste and waste exhumed from each of the three reactors after operation.

Waste characteristic Initial MSW MSW at end of reactor operation Reactor A Reactor B Reactor C Cellulose – C (%) 19.6 4.4 10.1 12.5 Hemicellulose – H (%) 5.4 1.6 4.0 4.3 Lignin – L (%) 28.8 43.5 30.6 33.1 [C + H]/L 0.87 0.14 0.46 0.51 Volatile solids (%) 52.9 25.6 56.7 53.8

Biochemical methane potential (mL-CH4/g-dry) 51.4 12.4 12.0 9.3 Back-calculated initial C (%)a NA 5.5 12.8 15.9 Back-calculated initial H (%)a NA 1.7 4.3 4.7

a Computed based on C and H required for cumulative methane yield during reactor operation (Table 2) plus C and H measured on exhumed waste at end of reactor operation. fresh leachate was initially collected from Deer Track Park Landfill et al. (2011b). Blended materials were hand squeezed through a such that the inoculums and all subsequent leachate additions for paint strainer bag to remove large solids. The strained liquid was the three reactors were from the same leachate source. Fresh filtered on a 0.2-lm pore filter and DNA was extracted similar to leachate as well as excess leachate generated from the reactors the leachate samples. DNA extractions from the four solid waste (i.e., >1 L) was stored in zero-head-space containers at 4 °C to min- samples representing the end-state analysis for Reactor A (Fig. 1) imize biological activity. were pooled for pyrosequencing. Leachate samples were collected weekly and measured for pH, oxidation–reduction potential (ORP), and chemical oxygen demand (COD). pH and ORP were measured on a Fisher Scientific 2.5.2. Quantitative polymerase chain reaction (Waltham, MA) AR60 with Microelectrodes, Inc. probes (Bedford, DNA samples were screened for the presence of Archaea and NH; MI-410 for pH and MI-800/410 for ORP). Chemical oxygen de- specific methanogens (orders: Methanomicrobiales, Methanobacte- mand was measured using a HACH (Loveland, CO) 0–1500 mg/L riales, Methanococcales, and Methanosarcinales, and families: Met- COD test kit. Standard COD curves were created with potassium hanosarcinaceae and Methanosaetaceae) using endpoint hydrogen phthalate solution. All absorbance measurements were polymerase chain reaction (PCR) and the primers and PCR condi- completed with a SpectronicÒ 20 Genesys™ spectrophotometer tions reported by Yu et al. (2006). Visual examination on ethi- (Sigma–Aldrich Co., St. Louis, MO) at 600 nm wavelength. dium-bromide stained 1% agarose gels was used to confirm presence of the amplified product. Quantitative 50-exonuclease PCR (qPCR) standards were constructed via cloning endpoint PCR 2.4. Gas production and chemistry products selected from samples showing positive bands using the TOPO-TA vector (Invitrogen™, Life Technologies Co., Grand Island, Gas volume was measured via water displacement in calibrated NY) following the manufacturer’s instructions. Cloned products 50-L (±0.5 L) or 10-L (±0.1 L) containers submerged in acidified used to construct qPCR standard curves were sequenced and com- water (pH 3.0). Samples for composition analysis were extracted pared to the NCBI nucleotide database using BLAST to confirm from gas collection bags and stored in evacuated glass vials prior appropriate phylogeny. DNA concentration was assessed using to analysis. Gas composition was assessed with a Shimadzu Gas the NanoDrop 1000 (Thermo Fisher Scientific, Waltham, MA). Copy Chromatograph (GC-2014) equipped with flame ionization and numbers of standards were calculated from the DNA concentration thermal conductivity detectors. Percentages of hydrogen (H2), using the known length of the vector and insert. nitrogen (N2), oxygen (O2), carbon dioxide (CO2), and methane Quantitative 50-exonuclease PCR was used to determine copy (CH4) were calculated with respect to standard gases (Scott Spe- numbers of methanogen 16S ribosomal DNA following methods cialty Gases, Plumsteadville, PA; AGA Gas, Inc., Maumee, OH). in Yu et al. (2006). qPCR was performed on an iCycler (Bio-Rad Lab- oratories, Inc., Hercules, CA) with initial denaturation for 15 min at 2.5. Microbial analysis 94 °C, followed by 45 1 min cycles that included 30 s denaturing at 94 °C and 30 s annealing/elongation at 60 °C. Reactions included 2.5.1. DNA extraction 12.5 ll of Bullseye Hot Start Taq 2X mix (MidSci, St. Louis, MO), 0.5 Leachate effluent samples were collected weekly from each units additional Bullseye Hot Start Taq Polymerase (final concen- reactor and analyzed individually (i.e., samples from separate reac- tration 3 U/25 ll), 1 ll of 25-mM MgCl2 (final concentration tors were not pooled). Samples were filtered on 0.2-lm pore filters 3.5 mM), 500-nM (final concentration) of forward and reverse pri- (Pall Gelman Co., Ann Arbor, MI) and frozen at 20 °C until DNA mer, and 200-nM (final concentration) of dual labeled probe. No extraction. DNA was extracted using a PowerSoil DNA extraction more than 5 ng of DNA was used as template. kit (Mo-Bio, Carlsbad, CA) following manufacturer’s instructions. All qPCR reactions had efficiencies between 90% and 110% and However, the initial vortexing step was replaced by beadbeating standard curves were linear with a coefficient of determination in a BioSpec (Bio Spec Products Inc., Bartlesville, OK) beadbeater (R2) P 0.99. Fluorescence cycle thresholds were determined auto- for 3.5 min. DNA concentrations were assessed on a Thermo Scien- matically with the iCycler software using the maximum slope tific ND-1000 (Thermo Fisher Scientific Inc., Waltham, MA) using method and were between 25 and 50 fluorescence units. Standard absorbance at 260 nm. curves were run in duplicate for each qPCR assay, using 10-fold Municipal solid waste samples (50 g wet mass) were exhumed dilutions of cloned standard DNA. The starting concentration of from Reactor A at the end of reactor operation (Day 181) at the four target was calculated from the number of copies in the qPCR reac- locations shown in Fig. 1 for 16S rRNA gene tag pyrosequencing. tion multiplied by the dilution factor of the PCR, and divided by the Each solid waste sample was blended for 60 s in a Waring indus- volume of leachate filtered (or in the case of solids, mass pro- trial blender (Waring Laboratory & Science, Torrington, CT) with cessed) for DNA extraction. Standard deviations of methanogen chilled phosphate buffer (pH 6.7) following the protocol in Staley abundances were determined by performing triplicate reactions. C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992 1985

2.6. 16S rRNA gene tag pyrosequencing 60 Day 41 - Initiate leachate 454 pyrosequening was conducted on the V6 hypervariable re- addition/recirculation gion of the 16S rRNA gene to assess bacterial community composition. PCR was carried out using the barcoded primers of Hamady et al. (2008) and 454 data was processed using recommendations 50 C)

of Huse et al. (2007). Sequences were analyzed with mothur v. o 1.27.0 (Schloss et al., 2009) using the standard ‘‘Schloss SOP’’. Briefly, flowgrams were quality-controlled and de-noised with the commands trim.flows, shhh.flows, trim.seqs; aligned against 40 the Silva reference database using align.seqs; and chimera-checked with the command chimera.uchime. The final sequence dataset was classified using the Silva taxonomy distributed with mothur, using the command classify.seqs. Operational taxonomic units Waste Temperature( (OTUs) were identified for the 0.03 distance level following con- 30 struction of a distance matrix using the commands dist.seqs and Reactor A cluster. Alpha diversity estimates were conducted using the rare- Reactor B faction.single and summary.single commands. Comparison of com- Reactor C munity composition was conducted using dist.shared and the Bray 20 Curtis dissimilarity index (DBC), which is calculated using the fol- 0 50 100 150 200 lowing equation: P Elapsed Time (d) PminðSAP;i; SB;iÞ DBC ¼ 1 2 ð1Þ Fig. 2. Average weekly waste temperature during reactor operation. SA;i þ SB;i where S is the number of individuals in the ith OTU of community A,i chemical parameters, and are consistent with previously described A and S is the number of individuals in the ith OTU of community B,i leachate chemistry behavior for anaerobic decomposition of MSW B. Effluent leachate samples from all reactors at peak methane gen- (e.g., Barlaz et al., 1989; Pohland and Kim, 1999; Laloui-Carpentier eration as well as leachate and solid waste samples from Reactor A et al., 2006). Stimulation of hydrolytic and fermentative bacteria at the end-state of the test were analyzed. These samples were se- with the onset of leachate addition (Day 41) produced an accumu- lected to compare microbial community composition between (1) lation of carboxylic acids, indicated by high COD concentrations leachates representative of different peak methane productions, and near or below neutral pH (Fig. 3). Subsequently, COD concen- (2) leachate and solid fractions at similar states of MSW decompo- trations decreased, pH increased, and ORP approached a minimum sition, and (3) leachate at peak methane production and end-state (Fig. 3); leachate chemistry trends that are indicative of a transition decomposition. to active methanogenesis. Following the removal of readily avail- Pyrosequencing was successful for effluent leachate samples able substrates in the leachate (80–100 d), COD remained below from Reactors B and C at peak methane flux and on end-state 10,000 mg/L and pH remained greater than 7.5 in all reactors for leachate and solid samples from Reactor A; all yielded >2000 se- the duration of the experiment. quences following quality control procedures. The leachate sample from Reactor A at peak methane yielded an unusually low number 3.1.3. Methane generation of sequences and was not included in subsequent analysis. Temporal relationships of methane flow rates and cumulative methane yield for the three reactors are shown in Fig. 4. Methane 3. Results composition of the biogas was similar in all reactors and averaged between 61% and 64% from 90 d to the end of the experiment, with 3.1. Bioreactor performance the balance being CO2. All three reactors exhibited higher initial methane flow rates characteristic of accelerated methane produc- 3.1.1. Waste temperature tion phases, followed by decreasing flow rates indicating a transi- Temporal trends of weekly average waste temperature for each tion to a decelerated methane production phases (Barlaz et al., reactor are shown in Fig. 2. Temperatures plotted in Fig. 2 were 1989). averaged from thermocouples installed within MSW specimens Cumulative methane yield, elapsed time for onset of methane (i.e., temperatures measured in the gravel are not included) production and peak methane flow rate, peak and average meth- (Fig. 1). Temperature control issues resulted in waste temperatures ane flow rates, and average waste temperature before and during in Reactor B varying from 22 to 54 °C between 29 and 60 d. During methane production for each reactor are summarized in Table 2. this period, waste temperatures were more stable in Reactor A Although the initial MSW composition was similar between all (27–34 °C) and Reactor C (27–39 °C). Average waste temperatures three reactors, cumulative methane yield ranged from 5.25 to for all reactors were relatively constant between 60 and 181 d: 18.7 L-CH4/kg-dry-MSW (Table 2, Fig. 4). These differences in 35.9 ± 3.3 °C in Reactor A, 39.2 ± 1.4 °C in Reactor B, and cumulative methane yield are attributed to variability in the initial 34.4 ± 2.2 °C in Reactor C. Although waste temperatures and tem- cellulose and hemicellulose compositions between the three reac- poral changes differed between reactors (Fig. 2), waste tempera- tors (described subsequently). The highest peak and average meth- tures between 60 and 181 d are representative of mesophilic ane flow rates were measured in Reactor C, which also generated environments. the largest cumulative methane yield (Table 2).

3.1.2. Leachate chemistry 3.1.4. Discussion Relationships of pH, COD, and ORP measured in the efﬂuent The elapsed time to methane production was shortest in Reac- leachate vs. time for all three reactors are shown in Fig. 3. Similar tor C, with measurable methane production commencing on Day temporal trends are observed in each reactor for all three leachate 48, 7 d after the ﬁrst 10-L leachate dose. Prior to initiation of 1986 C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992

(a) Reactor A Reactor B Reactor C Reactor A Reactor B Reactor C 8.0 0.8 20 Reactor C R1 R2 (L-CH Yield Methane Cummulative R3 7.5 0.6 15 Reactor B pH /kg-MSW/d) 7.0 4

0.4 10 Inoculum pH = 8.0 6.5 (b) Reactor A 50000 4 /kg-MSW) Inoculum COD = 2400 mg/L 0.2 5 40000 Methane Flow (L-CH Flow Rate Methane

30000

20000 0.0 0 40 80 120 160 200 Elapsed Time (d) 10000 Fig. 4. Temporal relationships of methane ﬂow rates and cumulative methane yield

Chemical (mg/L) Demand Oxygen Chemical 0 for all three reactors. Note: symbols represent methane ﬂow rates and lines (c) represent cumulative methane yield for each reactor. 100 lag-time for onset of methane production in Reactor B is also sup- 0 ported by lag-times in COD depletion and pH stabilization compared to Reactor C (Fig. 3). -100 Leachate chemistry trends for Reactor A suggest that the duration of active methanogenesis was shorter compared to Reactors B -200 and C. pH stabilized to above 7.4 (Fig. 3a), COD decreased to less than 10,000 mg/L (Fig. 3b), and ORP transitioned from negative -300 to positive (Fig. 3c) more rapidly in Reactor A compared to Reactors Inoculum ORP = -370 mV B and C. These transitions, combined with limited early methane

Oxidation Reduction Potential (mV) -400 generation, suggest a shorter methanogenic phase occurred in 40 80 120 160 200 Reactor A. However, the magnitude of COD concentration, which indicates the relative abundance of substrates for methanogenesis, Elapsed Time (d) was similar between all three reactors. Thus, while temporal Fig. 3. Temporal relationships of leachate (a) pH, (b) chemical oxygen demand, (c) trends in leachate chemistry are indicative of microbial activity and oxidation–reduction potential for all three reactors. First leachate addition on contributing to MSW decomposition, the overall magnitude of Day 41 and first sample collected on Day 48. the different chemical parameters (e.g., COD) do not appear to correlate to methane yield. methane production, the average waste temperature in Reactor C (Table 2) was near optimal for mesophilic waste decomposition 3.2. Microbial community and population (38–42 °C; Farquhar and Rovers, 1973; Barlaz et al., 1989) and more stable compared to the other two reactors (Fig. 2). This sta- 3.2.1. Bacteria ble, near optimal mesophilic temperature phase probably contrib- 16S rRNA gene tag pyrosequencing was conducted to assess uted to earlier methane generation in Reactor C by allowing rapid community diversity of leachate samples in Reactors B and C at establishment of a methanogenic community. peak methane production, and in Reactor A leachate and solids at Methane production in Reactors A and B was not observed until the end of the experiment. The total number of high-quality se- Day 75, 34 d after leachate addition began. The longer lag time for quences obtained and various measures of community diversity onset of methanogenesis in Reactor A was possibly a result of low- are presented in Table 3. Based on coverage estimates, more than er waste temperatures (Fig. 2) that retarded activity of microbial 90% of the predicted diversity was captured in the sequencing ef- communities. Decreases in methanogenic activity and methane fort. Interestingly, Reactor A end-state leachate had markedly high- generation have been reported with temperature decreases com- er alpha diversity (as measured by Simpson’s index) than any of parable to those measured in Reactor A (Fig. 2) following the initial the other samples. leachate dose (Hartz et al., 1982; Kettunen and Rintala, 1997). In The percent of sequences belonging to specific bacterial taxo- addition, Kettunen and Rintala (1997) report an increase in the nomic groups are summarized in Tables 4 and S1. Bacteria associ- lag time for methane generation from leachate for temperature ated with the phyla Bacteroidetes and Firmicutes constituted greater decreasing below 35 °C. The initial temperature fluctuations in than 70% of the sequences obtained from leachate in Reactors B Reactor B (Fig. 2) may have stagnated microbial development dur- and C at peak methane generation. Bacteroidetes and Firmicutes ing initial leachate recirculation (41–75 d), as temperatures fluctu- are common in anaerobic environments (e.g., anaerobic digester, ating between mesophilic and thermophilic ranges have been rumen), where they ferment complex polysaccharides such as cel- shown to inhibit waste decomposition (e.g., Pfeffer, 1974). The lulose and starch (e.g., Krause et al., 2008; Jaenicke et al., 2011). C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992 1987

Table 2 Summary of elapsed times for onset of methane generation and peak methane ﬂow rate, peak and average methane ﬂow rates, cumulative methane yield, and average waste temperature before and during methane generation for each reactor.

Characteristic Unit Reactor A Reactor B Reactor C

Elapsed time to CH4 generation d 75 75 48

Elapsed time to peak CH4 ﬂow rate d 91 99 86

Peak CH4 ﬂow rate L-CH4/kg-dry/d 0.53 0.43 0.69

Average CH4 ﬂow rate L-CH4/kg-dry/d 0.19 0.21 0.37

Cumulative CH4 yield L-CH4/kg-dry 5.27 14.25 18.7

Average waste temperature before CH4 generation °C 29.5 ± 2.3 40.6 ± 7.7 38.2 ± 2.3

Average waste temperature during CH4 generation °C 32.6 ± 0.8 38.3 ± 0.7 34.6 ± 2.4

Table 3 Sequencing results and alpha diversity estimates for leachate samples from Reactors B and C at peak methane generation and end-state leachate and solid samples from Reactor A based on 16S rRNA gene tag pyrosequencing.

Sample Number of sequencesa Coverageb (%) OTUsc Alpha diversityd Reactor B, peak methane 2170 90 331 5.38 Reactor C, peak methane 2541 91 353 5.18 Reactor A, end-state leachate 2032 90 316 21.18 Reactor A, end-state solids 2542 92 302 9.37

a Number of sequences remaining in the dataset following quality control. b Estimated Good’s coverage based on sequencing effort. c Number of operational taxonomic units (OTUs) identiﬁed at the 97% sequence identity cutoff. d Alpha diversity as measured by the inverse Simpson’s index. Higher values indicate higher diversity. In this case, Reactor A end-state leachate was the most even community.

Table 4 Percent of sequences belonging to speciﬁc phyla, orders, and families of Bacteria as determined by 454 tag pyrosequencing of 16S rRNA genes. Data are compiled for leachate samples from Reactors B and C at peak methane generation and end-state leachate and solid samples from Reactor A.a

Phylum Order Family Reactor B peak Reactor C peak Reactor A end-state Reactor A end-state methane (%) methane (%) leachate (%) solids (%) Bacteroidetes Bacteroidia Bacteroidales 10.8 9.5 4.0 3.4 Bacteroidetes Flavobacteria Flavobacteriales – – 1.4 – Bacteroidetes Sphingobacteria Sphingobacteriales 45.8 45.7 18.2 3.5 Chloroﬂexi Anaerolineae Anaerolineales ––– 1.9 Firmicutes Bacilli Bacillales ––– 5.0 Firmicutes Bacilli Lactobacillales 3.6 1.3 – – Firmicutes Clostridia Clostridiales 15.8 22.9 9.4 28.9 Firmicutes Clostridia Thermoanerobacterales – – 1.3 1.3 Firmicutes Mollicutes unclassiﬁed – 1.4 – – Lentisphaerae Lentisphaeria Victivallales 1.7 1.2 – – Proteobacteria Alphaproteobacteria Caulobacterales – – 4.8 – Proteobacteria Alphaproteobacteria Sphingomonadales – – 1.6 – Proteobacteria Betaproteobacteria Burkholderiales – – 9.2 – Proteobacteria Deltaproteobacteria Desulfovibrionales – – 1.3 – Proteobacteria Epsilonproteobacteria Campylobacterales 4.1 1.7 – – Proteobacteria Epsilonproteobacteria Nautiliales ––– 1.6 Proteobacteria Gammaproteobacteria Alteromonadales – – 8.9 – Proteobacteria Gammaproteobacteria Pseudomonadales – – 23.4 – Proteobacteria Gammaproteobacteria Xanthomonadales – – 2.4 – Spirochaetes Spirochaetes Candidatus_Cloacamonas – – 1.2 2.1 Spirochaetes Spirochaetes Spirochaetales ––– 4.4 Synergistetes Synergistia Synergistales 1.0 – 1.9 12.3 Thermotogae Thermotogae Thermotogales 11.0 10.1 3.0 30.7

a Only those phylogenetic groups representing greater than 1% of sequences in any one sample are presented here, and therefore percentages do not add to 100%. Table S1 in the supplementary online material contains a more complete list of taxa identiﬁed in the samples.

Bacteroidetes-like sequences are also commonly recovered from and belonged to the WCHB1-69 group within the Sphingobacteri- bioreactor landfill leachate and MSW (e.g., Huang et al., 2004; ales (Bacteroidetes phylum). The second most abundant OTU was Cardinali-Rezende et al., 2009). also found in all samples and was classified in the Petrotoga genus The extent of similarity in community composition among the within the Thermatogales (Thermotogae phylum). four samples was explored using the Bray Curtis metric (Table 5). Similarity in microbial community composition between Reac- The assemblages of bacterial operational taxonomic units (OTUs) tors B and C at peak methane generation suggests that temperature in the leachate of Reactors B and C at peak methane generation fluctuations prior to methanogenesis (Fig. 4) did not influence were approximately 70% similar by this metric, and were markedly microbial community composition. Reactor B experienced a pro- distinct from the assemblages in Reactor A leachate and solids at nounced increase in waste temperature prior to leachate addition, the end-state. The most abundant OTU across the whole dataset followed by fluctuating temperatures before equilibrating at (comprising 49% of all sequences) was found in all four samples approximately 39 °C and transitioning to methanogenesis (Day 1988 C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992

Table 5 Community similarity based on 97%-OTUs identiﬁed by 16S rRNA gene tag pyrosequencing, based on the Bray Curtis similarity index.

Sample Reactor B peak methane Reactor C peak methane Reactor A end-state leachate Reactor C peak methane 0.709 — — Reactor A end-state leachate 0.189 0.193 — Reactor A end-state solids 0.196 0.176 0.154

75). Reactor C experienced decreasing waste temperatures be- 3.2.2. Archaea and methanogens tween the onset of leachate addition (Fig. 2) and peak methane Temporal relationships of methanogen abundance (orders Met- generation (Day 86) from approximately 39 to 31 °C. While the hanomicrobiales and Methanobacteriales and family Methanosarcin- fluctuating waste temperatures in Reactor B are believed to have aceae) measured with qPCR on effluent leachate for each reactor prolonged the onset of methanogenesis by inhibiting growth, var- are shown in Fig. 5. Methane flow rates are reproduced as weekly iability in waste temperatures between Reactors B and C did not averages in Fig. 5 to identify relationships between methane influence the active microbial community composition after tem- generation and methanogen dynamics. Abundance of all three perature stabilization. The similar microbial communities were methanogens (Methanomicrobiales, Methanosarcinaceae, and attributed to similarity in the waste source and leachate inoculum, Methanobacteriales) increased between 48 and 97 d, concurrent combined with similar waste temperatures within the mesophilic with increasing methane flow rates (Fig. 5). Peak methane flow range during waste decomposition and microbial community rates approximately coincided with maximum methanogen abun- development. dance, and subsequent methanogen abundances remained ele- Bacterial community compositions determined for the end- vated for the duration of the experiment in all three reactors. state leachate and solid samples in Reactor A were not similar (Tables 4 and 5, and S1). Sequences retrieved from the solid fraction in Reactor A revealed a larger percentage of Clostridia com- (a) Msc MMB MBT CH Rate pared to Reactor A leachate as well as both Reactor B and 4 (L-CH Rate Flow Methane 109 0.8 Reactor C leachates (Table 3). Thermotogae, another anaerobic fer- Reactor A mentative organism, was also more abundant in the solids, accounting for >31% of the solids sequences (Table S1). The pres- 107 0.6 ence of Thermotagae at mesophilic temperatures confirms recent reports of abundant mesophilic members of a genus previously 105 0.4 considered to be predominantly hyperthermophilic (Nesbø et al., 2006). Thermotogae ferment complex polysaccharides,

3 4 /kg-dry/d) including cellulose, and produce hydrogen gas as a byproduct (16S copies/mL) 10 0.2

(Conners et al., 2006; Cardinali-Rezende et al., 2009). Thermotogae Methanogen Abundance and Clostridia may have been enriched in the solids fraction due 101 0.0 to a preference for growth attached to a solid polysaccharide sub- Methane Flow Rate (L-CH strate, such as cellulose. (b) 109 0.8 Proteobacteria accounted for approximately 55% of sequences in Reactor B the leachate, but less than 3% in solids, for Reactor A at the end- 7 state (Table S1). Members of the Proteobacteria are physiologically 10 0.6 very diverse and many can ferment simple polysaccharides or monosaccharaides (Krause et al., 2008; Jaenicke et al., 2011). These 105 0.4 bacteria were likely enriched in the leachate because they were involved in fermentation of soluble substrates such as simple sugars.

3 4 /kg-dry/d) The differences between bacterial communities in solids and leach- copies/mL) (16S 10 0.2 ate fractions of MSW agree with recent ﬁndings by Staley et al. Abundance Methanogen (2012), and further supports the assessment of both solids and 101 0.0 leachate to more broadly capture the bacteria community composition of MSW. (c) 109 0.8 Methane Flow Rate (L-CH A direct comparison between bacteria community composition Reactor C in the leachate fraction at peak methane generation and end-state 107 0.6 decomposition is not possible due to failure of the pyrosequencing reaction for Reactor A peak methane. However, assuming the bacterial communities in Reactors B and C are representative 105 0.4 of peak-methane leachate in Reactor A, the bacterial community shows a shift from a predominantly Bacteroidetes and Firmicutes 3 4 (16S copies/mL)

10 0.2 /kg-dry/d) community during peak methane production (Reactors B and C in Table 4)toaProteobacteria-dominated community during Methanogen Abundance end-state decomposition (Reactor A in Table 4). This change could 101 0.0 represent a depletion of complex polysaccharides, for instance as 40 80 120 160 200 cellulose is degraded and removed, and a shift towards degrada- Elapsed Time (d) tion of less complex breakdown products such as sugar monomers. This progression was previously described by Barlaz et al. Fig. 5. Temporal relationships of methanogen abundance in leachate and methane (1989), who measured a decrease in cellulolytic bacteria and an ﬂow rate for Reactor A (a), Reactor B (b), and Reactor C (c). Msc = Methanosarcin- aceae, MMB = Methanomicrobiales, and MBT = Methanobacteriales. Vertical dashed increase in acetogenic bacteria during decelerated methane lines indicate transitions from accelerated to decelerated methane generation production. phases. C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992 1989

This is consistent with previous work showing elevated methano- abundant in the leachate fraction during accelerated methane pro- gen abundance throughout the decelerated methane phase (Barlaz duction due to initial hydrolysis, fermentation, and acetogenesis of et al., 1989). The order Methanococcales and the family Methano- cellulose and hemicellulose in MSW (Barlaz et al., 1989). Following saetaceae were not detected by PCR in any leachate sample. the removal of readily available substrates in the leachate fraction, Although the PCR-based approaches to detecting 16S rRNA characterized by the transition from a high rate of COD removal to genes are powerful tools for studying microbial community com- a low rate of COD removal between 90 and 100 d (Fig. 3b), methane position, they do have limitations that should temper interpreta- generation rates decrease as waste decomposition transitions to a tion of the results. These approaches are inherently focused on decelerated methane production phase (Fig. 5). DNA and only provide a proxy for the abundance of different target Relationships between methane flow rate and methanogen phylotypes (groups). They do not directly measure activity of the abundance during accelerated and decelerated methane genera- detected organisms or the expression of genes involved in key tion phases are shown in Fig. 6. During accelerated methane gener- metabolisms. However, the presence and abundance of specific ation, methane flow rate vs. methanogen abundance data from all populations is assumed necessary to drive the measured processes reactors create a single relationship that indicates higher methane (i.e. hydrolysis, fermentation, volatile fatty acid oxidation, and flow rates correspond to higher methanogen abundance. In con- methanogenesis). In any case, technological advances allowing trast, there is no correspondence between methane flow rate and detection of specific mRNAs and whole transcriptomes may reveal methanogen abundance during the decelerated methane produc- new relationships in future studies. tion phase (Fig. 6b). In anaerobic decomposition, methanogenesis is rate-limiting when substrates are solubilized (e.g., during accelerated methane production) and hydrolysis is rate-limiting when 3.3. Methanogen population and methane generation substrates are complex solid organic matter (Nioke et al., 1985; Pavlostathis and Giraldo-Gomez, 1991; Vavilin et al., 1996). The 3.3.1. Implications on methane flow rates positive correspondence shown in Fig. 6a suggests methanogen Temporal relationships of methanogen abundance and methane growth drives methane generation when soluble substrates are flow rates shown in Fig. 5 indicate that the initial methanogen readily available (e.g., high COD concentration), and that higher growth phase coincided with the accelerated methane production methane flow rates are dependent on establishing of a larger meth- phase. Substrate (e.g., acetate) for methane generation is readily anogen population. Methane flow rates for all three reactors ranged between

approximately 0.01 and 0.3 L-CH4/kg-dry/d during decelerated (a) Reactor A Reactor B Reactor C methane production. However, methanogen abundance within 1 each reactor remained nearly consistent throughout the decelerated methane production phase, corresponding to the abundances achieved during the growth phase (Fig. 5). The continued detection of high methanogen abundance following peak methane genera-

/kg-dry/d) tion suggests that methanogens remain abundant in MSW, but at 4 0.1 low metabolic activity after the bulk of readily degradable organic matter have been consumed. Thus, the presence of methanogens may be applicable as an indicator of past or potential future methane production, but is not necessarily a direct indicator of current 0.01 methanogenic activity.

Accelerated Methane

Methane Flow Rate (L-CH Rate Flow Methane 20 Production Phase 0.001 4 5 6 7 8 9 (b) 10 10 10 10 10 10 1 15 /kg-dry) 4 /kg-dry/d) 4 0.1 10

0.01 5 Methanomicrobiales Reactor Methane Yield(L-CH Methane Reactor Decelerated Methane Methanobacteriales

Methane Flow Rate (L-CH Methanosarcinaceae Production Phase Total Methanogens 0.001 104 105 106 107 108 109 0 104 105 106 107 108 109 Total Methanogen Abundance (16S copies/mL) Methanogen Abundance (16S copies/mL) Fig. 6. Methane ﬂow rates vs. total methanogen abundance in each reactor during (a) accelerated and (b) decelerated methane production phases. Dashed lines in (a) Fig. 7. Relationship between cumulative methane yield and average methanogen capture the general trend in the data. abundance between 97 and 181 d of reactor operation. 1990 C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992

3.3.2. Implications on methane yield back-calculated from cumulative methane yield (Table 2) and end- The relationship between cumulative methane yield and meth- state C and H measured on exhumed waste (Table 1). The back-cal- anogen abundance for the three reactors is shown in Fig. 7. Average culated initial C and H (Table 1 and Fig. 8) indicate that the initial abundance of Methanomicrobiales, Methanosarcinaceae, and Met- available substrate to support microbial growth differed between hanobacteriales in Fig. 7 are representative of methanogen popula- reactors. These different initial C and H contributions are attrib- tions measured on the leachate fraction between 97 and 181 d (i.e., uted to differences in the material specific biodegradable constitu- established methanogen population following the accelerated ents comprising the paper and cardboard fraction (e.g., office methane generation phase, Fig. 5), and total methanogens are the paper, newsprint, corrugated cardboard), which was the primary sum of average individual methanogen abundances for each reac- biodegradable fraction for MSW used in this study. MSW compo- tor. Although similar methanogen communities were observed nents comprising the paper/cardboard fraction have varying rates among reactors, higher abundance of Methanomicrobiales and Met- of decomposition and cumulative methane yield (Owens and Chy- hanosarcinaceae correlate with higher cumulative methane yield noweth, 1993; Eleazer et al., 1997; Staley and Barlaz, 2009). Thus, (Fig. 7). The total methanogen abundance for the higher two meth- although the initial waste composition was consistent between the ane yields (Reactors B and C) is dominated by Methanomicrobiales three reactors, the chemical signature of the MSW constituents (Fig. 7). This suggests Methanomicrobiales influenced methane gen- likely varied and contributed to different initial substrate abun- eration to a greater extent than Methanosarcinaceae or Methano- dance to support microbial growth. bacteriales, and that the majority of methane was produced via The back-calculated C and H (Table 1) are believed to more ade- the hydrogenotrophic pathway. The presence of abundant hydro- quately represent the initial C and H available in each reactor as genotrophic methanogens is in agreement with previous work on compared to the C and H measured on the initial MSW (Table 1). MSW and other anaerobic digestion systems (e.g., McMahon The lack of trends in the C, H, [C + H]/L, VS, and BMP between et al., 2001; Chen et al., 2003; Krause et al., 2008; Rastogi et al., the three exhumed wastes and initial waste (Table 1) are likely 2008; Nayak et al., 2009; Jaenicke et al., 2011; Sasaki et al., 2011; due to inherent heterogeneity in chemical properties of compo- Zhu et al., 2011; Staley et al., 2012). Interestingly, Thermotogae nents contained in each biodegradable group (Bareither et al., and Clostridia, two of the most abundant phyla in the reactors, 2010b). Relative percentages of C and H contributing to a methane are both known to produce hydrogen during anaerobic degrada- yield from a given reactor (Table 2) were based on waste decompo- tion of cellulose and other polysaccharides (Krause et al., 2009; sition measured in a field-scale experiment on MSW from the same Cardinali-Rezende et al., 2009; Jaenicke et al., 2011). This suggests source and with the same initial composition as used in this study that there was likely syntrophic relationships between hydrogen (Bareither et al., 2012). Comparison of initial to final C + H in the producing bacteria and hydrogen consuming methanogens in the field-scale experiment revealed that 93% of the change between MSW. The Bacteroidetes, on the other hand, are acetogenic bacteria initial and final biochemical methane potential was attributed to and likely supported the growth of acetotrophic methanogens such C + H decomposition (Bareither et al., 2012). This observation is as Methanosarcinaceae. in agreement with Barlaz et al., (1990), who report >90% of meth- The relationship between methanogen abundance and initial ane yield in MSW originates from C and H. percent contribution of cellulose (C) plus hemicellulose (H) in each Statistically significant correlations for Methanomicrobiales, reactor is shown in Fig. 8. The initial C and H for each reactor were Methanosarcinaceae, and total methanogen abundance versus initial C + H are shown in Fig. 8 for trendlines with coefficients of determination (R2) > 0.90. These correlations are evidence that Methanomicrobiales Methanosarinaceae methanogen abundance in MSW is dependent on the available C Methanobacteriales Total Methanogens and H, which are the primary substrates that support microbial growth in MSW. Relationships between cellulose degrading organ- 109 isms such as Thermotogae or Clostridia, and hydrogenotrophic β·(C+H) Exponential regression lines: M = C·e methanogens, likely provide the means for C + H to indirectly sup- Methanogen C β R2 port methanogen growth and activity. Comparisons between envi- MMB 3824 0.4860 0.995 8 ronmental factors measured in the MSW (e.g., waste temperature 10 MBT 148700 0.0207 0.250 Msc 40060 0.1426 0.911 and leachate chemistry) provide insights on the establishment Total 20190 0.4006 0.985 and proliferation of microbial communities. However, availability and type of substrate eventually determined the microbial com- 107 munity that developed in the MSW.

4. Conclusions 106 Bacterial diversity and methanogen abundance were evaluated in this study in three laboratory reactors simulating bioreactor landﬁll conditions. Quantitative molecular tools (50-exonuclease 105 qPCR) and next-generation sequencing (454 pyrosequencing) were

Methanogen Abundance (16S copies/mL) used to investigate microbial communities during biodegradation of municipal solid waste (MSW). The following conclusions are drawn from this study: 104 0 5 10 15 20 25 Similar microbial communities (bacterial and archaeal) develop Back-Calculated Initial Cellulose + Hemicellulose (%) during the accelerated methane production phase in MSW at mesophilic temperatures independent of initial temperature Fig. 8. Relationship between methanogen abundance determined from qPCR and ﬂuctuations. back-calculated initial cellulose plus hemicellulose content of the municipal solid waste. Msc = Methanosarcinaceae, MMB = Methanomicrobiales, and MBT = Methano- Temporal transitions in methanogenic activity as well as the bacteriales. duration of methanogenesis can be linked to transitions in C.A. Bareither et al. / Waste Management 33 (2013) 1982–1992 1991

leachate pH, chemical oxygen demand, and oxidation–reduc- Cardinali-Rezende, J., Debarry, R.B., Colturato, L.F.D.B., Carneiro, E.V., Chartone- tion potential. However, chemical oxygen demand concentra- Souza, E., Nascimento, A.M.A., 2009. Molecular identification and dynamics of microbial communities in reactor treating organic household waste. Appl. tion is not an indicator of cumulative methane yield. This may Microbiol. Biotechnol. 84 (4), 777–789. well be because the concentration does not adequately reflect Chen, A., Udea, K., Sekiguchi, Y., Ohashi, A., Harada, H., 2003. Molecular detection the flux of carbon through the ecosystem. and direct enumeration of methanogenic Archaea and methanotrophic Bacteria in domestic solid waste landfill soils. Biotechnol. Lett. 25, 1563–1569. 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