Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell

Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell

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Accepted Manuscript Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell Jungyu Park, Beom Lee, Donjie Tian, Hangbae Jun PII: DOI: Reference:

S0960-8524(17)31569-9 http://dx.doi.org/10.1016/j.biortech.2017.09.021 BITE 18849

To appear in:

Bioresource Technology

Received Date: Revised Date: Accepted Date:

30 June 2017 31 August 2017 1 September 2017

Please cite this article as: Park, J., Lee, B., Tian, D., Jun, H., Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell, Bioresource Technology (2017), doi: http://dx.doi.org/10.1016/j.biortech.2017.09.021

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Bioelectrochemical enhancement of methane production from highly concentrated food waste in a combined anaerobic digester and microbial electrolysis cell Jungyu Parka, Beom Leea, Donjie Tianb, Hangbae Juna,*

a

Department of Environmental Engineering, Chungbuk National University, Cheongju,

361-763, Republic of Korea b

JEONGBONG CO., LTD., 69-4 Munhwa-dong, Cheongju, Republic of Korea

Email addresses: Jungyu Park: [email protected] Beom Lee: [email protected] Dongjie Tian: [email protected] Hangbae Jun: [email protected]

Abstract A microbial electrolysis cell (MEC) is a promising technology for enhancing biogas production from an anaerobic digestion (AD) reactor. In this study, the effects of the MEC on the rate of methane production from food waste were examined by comparing an AD reactor with an AD reactor combined with a MEC (AD + MEC). The use of the MEC accelerated methane production and stabilization via rapid organic oxidation and rapid methanogenesis. Over the total experimental period, the methane production rate and stabilization time of the AD + MEC reactor were approximately 1.7 and 4.0 times faster than those of the AD reactor. Interestingly however, at the final steady state, the *Corresponding author ([email protected])

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methane yields of both the reactors were similar to the theoretical maximum methane yield. Based on these results, the MEC did not increase the methane yield over the theoretical value, but accelerated methane production and stabilization by bioelectrochemical reactions.

Keywords: microbial electrolysis cell, anaerobic digestion, food waste, methane production, microbial community

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1. Introduction Anaerobic digestion (AD) produces methane gas by the biodegradation and reduction of highly concentrated organic waste (Guo et al., 2013). However, AD is affected by substrate characteristics, operation temperatures, pH, alkalinity, ammonium ions, C/N ratio, volatile fatty acids (VFAs), nutrients, organic loading rate (OLR), reactor type, and toxicity (Wang et al., 2016; Rajagopal et al., 2013; Hobbs et al., 2017). Therefore, AD reactors exhibit unstable methane production and organic degradation (Moset et al., 2014; Appels et al., 2008). In particular, highly concentrated organic matter, such as food waste, inhibits methane production and stabilization via accelerated VFA accumulation and a decrease in pH (Hobbs et al., 2017). Lu et al. (2008) and Adekunle and Okolie (2015) showed that, for highly concentrated or complex organic matter, such as food waste and livestock wastewater, byproducts (complex heterocyclic compounds) or non-degradable VFAs form during hydrolysis, thus making it the ratelimiting step. Chen et al. (2008) reported that the methane production by an AD reactor is mainly inhibited by a decrease in pH and VFA accumulation at start-up, and this inhibition continues until the steady state is reached. Latif et al. (2017) also reported that a low pH is associated with various issues, including acid requirements, VFA accumulation, loss in methane production, and the inhibition of methanogenesis. Owing to the decreased pH and the accumulation of VFAs at start-up, an AD reactor often requires 2–9 months to become stabilized (Lauwers et al., 1990). Various researchers have investigated ways to improve the methane production rate and stabilization time. For example, one traditional method involves phase separation between acidogenesis (or hydrolysis) and methanogenesis to minimize interspecific competition and to increase the reaction rate at the rate-limiting step.

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Gough et al. (2013) reported that the methane production rate could be increased using separation by thermophilic acidogenesis at 55C and mesophilic methanogenesis at 35C; Lopez et al. (2014) obtained similar results. The recently developed microbial electrolysis cell (MEC), a microbial electrochemical technology, uses bioelectrochemical reactions to improve biogas production in an AD reactor by the rapid degradation of highly concentrated organic wastes, VFAs, toxic materials, and non-degradable organic matter (Zhang and Angelidaki, 2014). MECs supply a low voltage (0.2–0.8 V) to the AD reactor for bioelectrochemical reactions, in which exoelectrogenic bacteria decompose organic matter and release electrons at the anode (Eq. (1)). These electrons then move to the cathode in a closed circuit and are consumed, thereby producing CH4 (Eq. (2)) and H2 (Eq. (3)) (Logan et al., 2008). Anode: C2H4O2 + 2H2O → 2CO2 + 8H+ + 8e-

E0 = −0.28 V vs. SHE

(1)

Cathode: CO2 + 8H+ + 8e− → CH4 + 2H2O

E0 = −0.24 V vs. SHE

(2)

2H+ + 2e− → H2

E0 = −0.42 V vs. SHE

(3)

When a MEC was used with a single AD reactor, there was rapid degradation of not only the highly concentrated organic matter, but also the VFAs, toxic materials, and non-degradable matter (Zhang et al., 2013). The microbial activities and the rate of methane production were also increased by the bioelectrochemical reactions (Ding et al., 2016). Several studies have confirmed that the MEC results in greater methane production than that of an AD reactor; methane yields achieved using a MEC are 0.31– 0.41 L-CH4/g-COD (chemical oxygen demand), which is close to the theoretical maximum methane yield at 35C (Bo et al., 2014; Tartakovsky et al., 2011; Xafenias and Mapelli, 2014; Yin et al., 2016). 4

These results indicate that bioelectrochemical reactions increase methane production by improving microbial activities and the efficiency of the removal of organic matter, including VFAs. Previously, Zhao et al. (2014) used a bioelectrochemical system to resolve a high OLR and found that methane production increased by the bioelectrochemical activation of acetoclastic and hydrogenotrophic methanogenesis, without a decrease in pH or VFA accumulation. Gajaraj et al. (2017) reported that MEC-assisted AD systems of 0.3 V and 0.6 V increase the yields of methane from glucose degradation by 9.4 ± 0.4% and 9.4 ± 0.5%, respectively, compared with the yield for an AD reactor. However, these experiments were conducted in small-scale reactors using a low concentration of synthetic substrate, such as acetate and glucose. More recently, Cerrillo et al. (2016) used an AD reactor coupled with a MEC to overcome the organic and nitrogen overload from pig slurry and reported that the AD + MEC combined system was a promising strategy for stabilization against organic and nitrogen overloads. This research was notable as it utilized pig slurry, but the working volume of the reactors was still less than 4 L. Dang et al. (2016) reported that an AD + MEC reactor with a carbon-based electrode could enhance methane production from dog food waste at a higher OLR than that of an AD reactor. However, few studies have examined methane production from municipal food waste using bench-scale AD and AD + MEC reactors to verify the effectiveness of MEC. Therefore, in this study, the effect of a MEC on the methane production rate and stabilization time from highly concentrated food waste was investigated in relatively large-scale AD and AD + MEC reactors (working volume: 20 L). Based on the study results, the roles of MEC in the AD reactor are discussed, with a focus on the methane

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production rate and stabilization time of highly concentrated food waste.

2. Materials and Methods 2.1 Characteristics of the substrate Food waste sampled from a food waste treatment plant (FWTP) in the Republic of Korea was used as the substrate; the food waste was pre-treated by filtration using a 150-m sieve. The pre-treated food waste diluted to 60.3 ± 2.1 g-TCOD (total chemical oxygen demand)/L was injected into each reactor once a day. The total solid (TS), total volatile solid (TVS), total nitrogen (TN), and pH of the feed provided to each reactor were 5.2 ± 0.4%, 3.9 ± 0.3%, 1.2 ± 0.2 g/L, and 5.2 ± 0.7, respectively.

2.2 Setup of the reactors and electrodes To assess the effects of a MEC on the methane production rate in an AD reactor, experiments were conducted in two sequencing batch reactors (SBR). One reactor was a typical mesophilic single AD reactor and the other was a mesophilic single AD + MEC reactor. The seeding sludge was obtained from an AD reactor of the FWTP and was inoculated to each reactor. Figure 1(a) shows the configuration of the two reactors, comprising an acrylic cylindrical structure (280 mm diameter × 410 mm height). The total volume of each reactor was 25 L and the working volume was 20 L. The AD + MEC reactor contained six sets of electrodes of 150 mm in width and 300 mm in height; the electrodes were positioned vertically to the rim of the cylindrical reactor. Each electrode was composed of graphite carbon mesh coated with Ni to increase electrical conductivity. For the cathode, a complex metal catalyst solution was prepared by dissolving 30.125 g of MnSO4·H2O, 19.75 g of KMnO4, 0.5684 g of iron

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phthalocyanine (FePc), and 0.5761 g of copper phthalocyanine (CuPc) in 1 L of distilled water and stirring for 2 h. The prepared solution was heated for 1 min and 30 s using a microwave and cooled for 60 s. The same procedure was repeated 5 times, and the solution was subsequently fixed on the graphite carbon mesh as described previously (Song et al., 2014). To minimize the internal resistance and to prevent contact between the electrodes, which were less than 3 mm apart, a piece of non-woven fabric was placed between the anode and cathode (Figure 1(b)). Titanium wires were used to complete the external circuit and to connect the anodes and cathodes. Each reactor had an agitator to maintain homogeneous conditions and valves to inject the substrate and collect the biogas.

2.3 Operational conditions The two reactors were operated for 12 months with an OLR of 3.0 kg-COD/m3d (1 L of food waste diluted to 60.3 ± 2.1 g-TCOD/L was injected into the reactors once a day). The hydraulic retention time (HRT) was 20 d, and the operation reactor type was a SBR. A direct current power supply (DP30-05TP; TOYOTECH, Fremont, CA, USA) was used, with the voltage for the AD + MEC reactor set at 0.3 V. The two reactors were operated in a thermostatic room at 35C. The operational conditions of the reactors are described in Table 1.

2.4 Pyrosequencing and analyses 2.4.1 Sludge sampling and DNA extraction The sludge was sampled from the bulk of the AD and AD + MEC reactors at an intermediate steady state (200 d). DNA was extracted from 200 μL of samples using the 7

FastDNA SPIN Kit for Soil (MP Biomedical, LLC, Santa Ana, CA, USA) according to the manufacturer’s instructions.

2.4.2 Polymerase chain reaction (PCR) amplification and pyrosequencing The extracted DNA was amplified by primers targeting the V1 to V3 regions of the 16S rRNA gene using a previously described method (Hur et al., 2011). DNA was sequenced by Chunlab Inc. (Seoul, Korea) using a Roche/454 GS Junior system according to the manufacturer’s instructions. The processing of pyrosequencing data of 16S rRNA gene sequences was performed as previously described (Hur et al., 2011). Amplification was performed under the following conditions: initial denaturation at 95°C for 5 min, followed by 30 cycles of denaturation at 95°C for 30 s, primer annealing at 55°C for 30 s, and extension at 72°C for 30 s, with a final elongation at 72°C for 5 min. The amplified products were purified using the QIAquick PCR Purification Kit (Qiagen, Valencia, CA, USA). Equal concentrations of purified products were pooled and the short fragments (non-target products) were removed using the Ampure Beads Kit (Agencourt Bioscience, Beverly, MA, USA). The quality and product size were assessed using a Bioanalyzer 2100 (Agilent, Palo Alto, CA, USA) with a DNA 7500 chip. Mixed amplicons were subjected to emulsion PCR and then deposited on Picotiter plates.

2.4.3 Pyrosequencing data analysis The reads obtained from different samples were sorted based on the unique barcodes of each PCR product. The sequences of the barcode, linker, and primers were removed from the original sequencing reads. Any reads containing two or more

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ambiguous nucleotides or a low-quality score (average score < 25) or reads shorter than 300 bp were discarded. Potential chimera sequences were detected using the Bellerophone method, which involves the comparison of BLASTN search results between the forward and reverse sequence halves (Hur et al., 2011). Chimeric sequences were detected using UCHIME (http://www.drive5.com/uchime/) and the EzTaxon-e database (http://eztaxon-e.ezbiocloud.net) was used to assign taxa to pyrosequencing reads (Yi et al., 2014). The richness and diversity of samples were determined using the Chao1 estimator and Shannon diversity index. A 3% cutoff was used for OTU assignment. Random subsampling was conducted to equalize the read size of samples for comparative analyses. The overall phylogenetic distance among communities was estimated using Fast UniFrac and visualized using a principal coordinate analysis. To compare operational taxonomic units (OTUs) between samples, shared OTUs were obtained using a XOR analysis implemented in the CLcommunity program (Chunlab Inc., Seoul, Korea).

2.5 Analysis The TCOD was analyzed using the closed reflux and colorimetric chrome methods after solid-liquid separation using a centrifuge (MF80; Hanil, Seoul, Korea). The VFAs were analyzed by liquid chromatography (SDV50A; Younglin, Anyang, Korea) with an absorbance detector (UV725S) installed after filtering with a 0.45-µm membrane filter (Lot No. 40606100; Advantec, Durham, NC, USA). The biogas produced was collected in a Tedlar bag (50 L) and was quantified in a thermostatic room using the water substitution method. The composition of the biogas produced from the reactors was detected using a gas chromatographer (GOW-MAC, Series 580)

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attached to a thermal conductivity detector. Ultra-pure helium was used as a carrier gas for the analysis and its flow velocity was fixed at 15 mL/min with temperatures of the column, injector, and detector fixed at 50°C, 80°C, and 90°C, respectively. The pH was measured using a pH meter (Orion 420A+; Thermo Fisher Scientific, Waltham, MA, USA) and alkalinity, TS, TVS, and TN were analyzed using standard methods (Eaton et al., 1995).

3. Results and Discussion 3.1 TCOD removal and methane production The properties of the AD and AD + MEC reactors are described in Table 2. Experimental periods were divided into three stages based on trends in methane production from an AD reactor: start-up (1–69 d), intermediate steady state (70–289 d), and final steady state (290–365 d). In the start-up stage of the AD reactor, ammonium ion and VFA concentrations accumulated to 2.2 ± 0.4 and 6.2 ± 1.7 g/L, respectively, and the pH was 6.7 ± 0.6. Concurrently, the VFA/alkalinity ratio of the AD reactor was 1.2 ± 0.8, which is approximately four times higher than the optimum ratio for stabilizing methane production, as reported by Kardos et al. (2011). For the AD + MEC reactor, ammonium ion and VFA concentrations did not accumulate in the start-up stage. Furthermore, pH and alkalinity were 7.42 ± 0.18 and 6.885 ± 0.774 g/L, respectively. Thus, during start-up, the AD reactor was limited by accumulated VFAs and ammonium ions, as well as a decreased pH. In contrast, in the AD + MEC reactor, these inhibitory substances did not accumulate. The TCOD of the effluent and the amount of methane gas produced by the AD and AD + MEC reactors are shown in Figure 2. The TCOD of the influent during the

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total experimental period was 60.3 ± 2.1 g/L. During the start-up, for the AD reactor, the TCOD of the effluent was 43.5 ± 6.3 g/L, although methane was not produced during this stage. This was because the pH decreased to 6.0 and the concentration of VFAs increased to 80 g/L (Table 2), which inhibited methane production; instead, methane was produced after a period of 70 d. The TCOD of the effluent and the daily methane produced during the intermediate steady state was 24.6 ± 3.9 g/L and 7.2 ± 1.9 L/d, respectively. In the final steady state, the TCOD of the effluent (10.9± 1.2 g/L) and the daily methane produced (16.5 ± 1.7 L/d) were stabilized. For the AD + MEC reactor, during the start-up, the TCOD of the effluent was 21.0 ± 2.4 g/L, which is 2-times lower than that of the AD reactor. The methane production increased gradually from approximately 3.5 to 10.3 L/d. In the intermediate steady state, the TCOD of the effluent decreased to 17.1 ± 1.8 g/L and the daily methane production increased to 12.1 ± 2.2 L/d. In the final steady state, the TCOD of the effluent and the daily methane production were stabilized at 10.1 ± 1.1 g/L and 17.0 ± 1.6 L/d, respectively. Rapid stabilization of methane production was observed from the start-up to the final steady state, without a decrease in pH or an accumulation of VFAs in this case. Previously, Zhao et al. (2014) reported that reported that a MEC affected the activation of acetoclastic methanogenesis and hydrogenotrophic methanogenesis with a high VFA removal efficiency and without a decrease in pH. Various studies have reported that the use of a MEC not only increases the removal efficiency of high-concentration organic waste, VFAs, toxic matter, and nonbiodegradable matter, but also enables methane production from hydrogen ions via bioelectrochemical reactions (Yin et al., 2015; Feng et al., 2015; Xafenias et al., 2014). Likewise, in this study, the COD removal efficiency and the methane production rate

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were enhanced by the MEC. Thus, the AD + MEC reactor had a higher methane production rate than that of the AD reactor because the inhibitory factors for methanogenesis were prevented during start-up by the bioelectrochemical reactions. These findings indicate that a MEC increased the methane production rate, stabilization, and organic removal.

3.2 Methane production rates The methane production rates were plotted using the linear trend line tool of Microsoft Excel for the three experimental stages (Figure 3). The methane production rate of the AD + MEC reactor was 6.9 ± 3.4 L/d during start-up; no methane was produced in the AD reactor during this time. During the intermediate steady state, the methane production rates of the AD and AD + MEC reactors were 7.2 ± 2.9 and 12.1 ± 2.2 L/d, respectively, and those in the final steady state were nearly similar for the two reactors (AD: 16.5 ± 1.7 L/d, AD + MEC: 17.0 ± 1.6 L/d). The average methane production rate of the AD + MEC reactor during the total experimental period was approximately 1.7 times higher than that of the AD reactor. To summarize, the methane production rates of the AD and AD + MEC reactors in the final steady state were nearly equal, although the methane production rate of the AD + MEC reactor was accelerated via bioelectrochemical reactions. Previously, Zhang et al. (2013) reported that a MEC maximizes the COD removal efficiency and the amount of methane production. However, in this study, the MEC did not cause more methane to be produced at the final steady state, but rather accelerated the methane production rate and stabilization by eliminating factors inhibiting methane production via bioelectrochemical reactions. To produce more methane, it would be necessary to

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intentionally add an electron donor, such as hydrogen gas, which could be produced by water electrolysis at a high voltage (up to 1.2 V). In the current study, however, no electron donor was added and the voltage was maintained at 0.3 V. This is clearly reflected in the results for the methane yields of the AD reactor and the AD + MEC reactor.

3.3 Methane yields The theoretical maximum methane yield at standard temperature and pressure (0C and 1 atm) based on the CODrem was 0.35 L-CH4/g-CODrem. Because the biogas from AD and AD + MEC reactors was sampled in mesophilic conditions (35C), the actual methane yield at standard temperature and pressure was calculated using the ideal gas equation. The actual methane yields in the three experimental stages are shown in Figure 4. For the AD and AD + MEC reactors, the yields in the start-up stage were 0 and 0.21 ± 0.09 L-CH4/g-CODrem, respectively. During start-up, the lack of methane production from the AD reactor was considered to result from inhibition of the activation of methanogenesis by accumulated VFAs and a decreased pH. Conversely, in the case of the AD + MEC reactor, the methane yield increased rapidly as a result of rapid VFA removal and the conversion of hydrogen ions to methane via bioelectrochemical reactions. In the intermediate steady state, the actual methane yields of the AD and AD + MEC reactors were 0.28 ± 0.05 and 0.33 ± 0.02 L-CH4/g-CODrem, respectively. The methane yield of the AD + MEC reactor was close to the theoretical maximum methane yield after 70 d. In contrast, the actual methane yield of the AD reactor was close to the theoretical methane yield after 290 d, indicating that the AD reactor required a period for stabilization 4-times longer than that of the AD + MEC 13

reactor. In the final steady state, the actual methane yields of the AD and AD + MEC reactors were 0.33 ± 0.02 and 0.34 ± 0.02 L-CH4/g-CODrem, respectively, which were both similar to the theoretical maximum methane yield. These results may be interpreted to indicate that the MEC does not overcome the theoretical maximum methane yield based on removed COD, but accelerated the methane production rate and stabilization by means of rapid COD (including VFA) removal and methane production via bioelectrochemical reactions, when no additional electron donor is present. Thus, the MEC affects the methane production rate and stabilization of an AD via bioelectrochemical reactions.

3.4 Effects of pH and VFAs on methane production In AD reactors, inhibitory effects of a high concentration of VFAs have been reported, and a decrease in pH resulting from VFA accumulation is considered the main cause of the inhibition (Chen et al., 2008; Rajagopal et al., 2013; Siegert and Banks, 2005; Rajagopal et al., 2013). Therefore, in the present study, the effects of pH and VFAs on methane production in AD and AD + MEC were analyzed (Figure 5). In the AD reactor, no methane was produced for a hydrogen concentration up to 5E-8 mol/L (less than pH 7.3), indicating that methane production is sensitive to a decrease in pH (Appels et al., 2008; Hwang et al., 2004; Jung et al., 2000). For the AD + MEC reactor, approximately 12.2 ± 0.8 L/d methane gas was produced for up to 5.E-8 mol/L (less than pH 7.3) hydrogen ions. These results showed that the hydrogen ions produced from bioanodic oxidation were rapidly converted to methane in the AD + MEC. Thus, the MEC was able to resolve the inhibition caused by a decrease in pH. As no methane was produced in the AD reactor during start-up, the VFAs accumulated up to 6.207 g/L,

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which inhibited methanogenesis. However, in the AD + MEC reactor, methane was produced without highly concentrated VFA accumulation. Thus, the MEC could resolve the main factors associated with inhibition by rapidly converting the hydrogen ions and VFAs to methane via bioelectrochemical reactions during start-up with a highly concentrated food waste. These results confirm previous findings reported by Zhao et al. (2014), indicating that the inhibitory factors were negated by bioanodic oxidation and biocathodic methanogenesis.

3.5 Microbial community analysis Most previous MEC studies have focused on the microbes living on the surface of the electrodes and on the mechanisms of electron transfer between electrodes and the microbial cell (Choi and Sang, 2016; Omidi and Sathasivan, 2013). Recently, however, some researchers have reported that the bulk sludge of the AD + MEC reactor can enhance methane production to a greater degree than obtained by electrode surface microbes (Feng et al., 2016; Zhao et al., 2014). Therefore, in the present study, the bulk sludge of the reactors in the intermediate steady state (200 d) was sampled for comparison. In the AD and AD + MEC reactors, the community structures of bacteria and archaea during the intermediate steady state were analyzed. Clear differences in AC160630_g and FN436026_g were observed between the AD and AD + MEC reactors. However, as AC160630_g and FN436026_g have not yet been identified, further analyses conducted at the class level are discussed in another paper (Lee et al., 2017). The class Clostridia was dominant in both the AD and the AD + MEC reactors; members of this class are exoelectrogenic bacteria that can decompose organic matter

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and transform it into acetate (Feng et al., 2015; Jiang et al., 2016). Previously, Xafenias et al. [46] detected members of the class Clostridia in both the starting sludge and the MEC sludge and showed that they contribute to the improvement in COD removal efficiency in the bulk sludge and particularly on the surface of the electrode in the MEC. Another dominant class found in the two reactors was Bacteroidia; members of this class hydrolyze proteins and transform the amino acids generated in the process into acetate (Ariesyady et al., 2007; Feng et al., 2015; Zuo et al., 2008). Although no differences were observed with regard to the dominant species between the AD and AD + MEC reactors, the class Clostridia was more dominant than Bacteroidia and the proportion of members of the class Clostridia in the AD + MEC reactor was 10% higher than that in the AD reactor. This difference may be attributed to the increased proportion of exoelectrogenic bacteria in the bulk of the AD + MEC reactor via electrochemical reactions, resulting in a high efficiency of organic matter removal. When members of the domain archaea were analyzed, Methanobacterium was the dominant genus in the AD reactor, whereas Methanosarcina was the dominant genus in the AD + MEC reactor. Members of the genus Methanobacterium exist, in general, in the AD reactor and in biocathodes in the MEC (Egli et al., 1987), and the genus Methanosarcina is known to include members that can utilize both electrons and hydrogen ions for methane production, even in extreme environments (Galagan et al., 2002; Yin et al., 2016). For a more accurate assessment, species-level archaeal communities were analyzed, and these findings are illustrated in Figure 6 (c) and (d). The abundances of Methanobacterium beijingense and M. petrolearium, which were the dominant species in the AD reactor, were drastically lower in the AD + MEC reactor,

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and Methanosarcina thermophile and Methanobacterium formicicum were distributed at low frequencies in the AD reactor, but at higher frequencies in the AD + MEC reactor. M. beijingense and M. petrolearium, which were the dominant species in the AD reactor, are hydrogenotrophic methanogens that produce methane using certain VFAs, hydrogen, and carbon dioxide (Ma et al., 2005; Maus et al., 2013; Mori and Harayama, 2011). Methanosarcina thermophile, which was the dominant species in the AD + MEC reactor, performs acetoclastic methanogenesis to convert acetate, methanol, and other substances into methane, and Methanobacterium formicicum performs hydrogenotrophic methanogenesis to produce methane using formate, hydrogen, and carbon dioxide (Maus et al., 2013; Fournier and Gogarten, 2008; van Bruggen et al., 1984; Zinder et al., 1985). These results are similar to those reported by Feng et al. (2015), who showed that Methanosarcina is enriched in an AD + MEC reactor with 0.3 V and a graphite carbon cathode. Thus, it can be concluded that more organic matter was removed in the AD + MEC reactor than in the AD reactor owing to the higher proportion of exoelectrogenic bacteria and the rapid conversion of hydrogen ions and carbon dioxide generated during decomposition into methane by hydrogenotrophic methanogenesis. Furthermore, acetoclastic methanogenesis, which did not occur in the AD reactor owing to the decreased pH and accumulated VFAs, was activated by the simultaneous prevention of inhibitory factors in the AD + MEC reactor.

4. Conclusions The methane production rate and stabilization time of the AD + MEC reactor were approximately 1.7 and 4.0 times faster than those in the AD reactor. The MEC

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could resolve the main factors inhibiting methanogenesis by rapid conversion of hydrogen ions and VFAs to methane via bioelectrochemical reactions during the startup. Interestingly, however, at the final steady state, the methane production rates of both reactors (i.e., AD and AD + MEC) were similar. Based on these results, the MEC did not overcome the theoretical methane yield, but accelerated methane production and stabilization by increased exoelectrogenic bacteria and acetoclastic methanogen communities.

Acknowledgements None

Funding: This work was supported by the Human Resource Training Program for Regional Innovation and Creativity through the Ministry of Education and National Research Foundation of Korea [grant number NRF-2015H1C1A1035673]. The sources had no role in the collection, analysis, and interpretation of data; in the writing of the report; and in the decision to submit the article for publication.

Declaration of Interest There are no conflicts of interest to be declared.

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References 1. Adekunle, K.F., Okolie, J.A., 2015. A review of biochemical process of anaerobic digestion. Adv. Biosci. Biotechnol. 6, 205–212. 2. Appels, L., Baeyens, J., Degrève, J., Dewil, R., 2008. Principles and potential of the anaerobic digestion of waste-activated sludge. Prog. Energy Combust. Sci. 34, 755–781. 3. Ariesyady, H.D., Ito, T., Okabe, S., 2007. Functional bacterial and archaeal community structures of major trophic groups in a full-scale anaerobic sludge digester. Water Res. 41, 1554–1568. 4. Bo, T., Zhu, X., Zhang, L., Tao, Y., He, X., Li, D., Yan, Z., 2014. A new upgraded biogas production process: Coupling microbial electrolysis cell and anaerobic digestion in single-chamber, barrel-shape stainless steel reactor. Electrochem. Commun. 45, 67–70. 5. Cerrillo, M., Viñas, M., Bonmatí, A., 2016. Overcoming organic and nitrogen overload in thermophilic anaerobic digestion of pig slurry by coupling a microbial electrolysis cell. Bioresour. Technol. 216, 362–372. 6. Chen, Y., Cheng, J.J., Creamer, K.S., 2008. Inhibition of anaerobic digestion process: A review. Bioresour. Technol. 99, 4044–4064. 7. Choi, O.K., Sang, B.I., 2016. Extracellular electron transfer from cathode to microbes: Application for biofuel production. Biotechnol. Biofuels. 9, 11. 8. Dang, Y., Holmes, D.W., Zhao, Z., Woodard, T.L., Zhang, Y., Sun, D., Wang, L.-Y., Nevin, P.N., Lovely, D.R., 2016. Enhancing anaerobic digestion of complex organic waste with carbon-based conductive materials. Bioresour. Technol. 220, 516–522.

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9. Ding, A., Yang, Y., Sun, G., Wu, D., 2016. Impact of applied voltage on methane generation and microbial activities in an anaerobic microbial electrolysis cell (MEC). Chem. Eng. J. 283, 260–265. 10. Eaton, A.D., Clesceri, L.S., Greenberg, A.E., Franson, M.A.H., 1995. Standard Methods for the Examination of Water and Wastewater, nineteenth ed. APHA. 11. Egli, C., Scholtz, R., Cook, A.M., Leisinger, T., 1987. Anaerobic dechlorination of tetrachloromethane and 1,2-dichloroethane to degradable products by pure cultures of Desulfobacterium sp. and Methanobacterium sp. Microbiol. Lett. 43, 257–261. 12. Feng, Q., Song, Y.C., Bae, B.U., 2016. Influence of applied voltage on the performance of bioelectrochemical anaerobic digestion of sewage sludge and planktonic microbial communities at ambient temperature. Bioresour. Technol. 220, 500–508. 13. Feng, Y., Zhang, Y., Chen, S., Quan, X., 2015. Enhanced production of methane from waste activated sludge by the combination of high-solid anaerobic digestion and microbial electrolysis cell with iron–graphite electrode. Chem. Eng. J. 259, 787–794. 14. Fournier, G.P., Gogarten, J.P., 2008. Evolution of acetoclastic methanogenesis in Methanosarcina via horizontal gene transfer from cellulolytic Clostridia. J. Bacteriol. 190, 1124–1127. 15. Gajaraj, S., Huang, Y., Zheng, P., Hu, Z., 2017. Methane production improvement and associated methanogenic assemblages in bioelectrochemically assisted anaerobic digestion. Biochem. Eng. J. 117, 105–112. 16. Galagan, J.E., Nusbaum, C., Roy, A., Endrizzi, M.G., MacDonald, P., FitzHugh,

20

W., Calvo, S., Engels, R., Smirnov, S., Atnoor, D., Brown, A., Allen, N., Naylor, J., Stange-Thomann, N., DeArellano, K., Johnson, R., Linton, L., McEwan, P., McKernan, K., Talamas, J., Tirrell, A., Ye, W., Zimmer, A., Barber, R.D., Cann, I., Graham, D.E., Grahame, D.A., Guss, A.M., Hedderich, R., Ingram-Smith, C., Kuettner, H.C., Krzycki, J.A., Leigh, J.A., Li, W., Liu, J., Mukhopadhyay, B., Reeve, J.N., Smith, K., Springer, T.A., Umayam, L.A., White, O., White, R.H., Conway de Macario, E., Ferry, J.G., Jarrell, K.F., Jing, H., Macario, A.J., Paulsen, I., Pritchett, M., Sowers, K.R., Swanson, R.V., Zinder, S.H., Lander, E., Metcalf, W.W., Birren, B., 2002. The genome of M. acetivorans reveals extensive metabolic and physiological diversity. Genome Res. 12, 532–542. 17. Gough, H.L., Nelsen, D., Muller, C., Ferguson, J., 2013. Enhanced methane generation during theromophilic co-digestion of confectionary waste and greasetrap fats and oils with municipal wastewater sludge. Water Environ. Res. 85, 175–183. 18. Guo, X., Liu, J., Xiao, B., 2013. Bioelectrochemical enhancement of hydrogen and methane production from the anaerobic digestion of sewage sludge in single-chamber membrane-free microbial electrolysis cells. Int. J. Hydrogen Energy. 38, 1342–1347. 19. Hobbs, S.R., Landis, A.E., Rittmann, B.E., Young, M.N., Parameswaran, P., 2017. Enhancing anaerobic digestion of food waste through biochemical methane potential assays at different substrate: inoculum ratios. Waste Manag. 20. Hur, M., Kim, Y., Song, H.R., Kim, J.M., Choi, Y.I., Yi, H., 2011. Effect of genetically modified poplars on soil microbial communities during the

21

phytoremediation of waste mine tailings. Appl. Environ. Microbiol. 77, 7611– 7619. 21. Hwang, M.H., Jang, N.J., Hyun, S.H., Kim, I.S., 2004. Anaerobic bio-hydrogen production from ethanol fermentation: The role of pH. J. Biotechnol. 111, 297– 309. 22. Jiang, Y.B., Zhong, W.H., Han, C., Deng, H., 2016. Characterization of electricity generated by soil in microbial fuel cells and the isolation of soil source exoelectrogenic bacteria. Front. Microbiol. 7, 1776. 23. Jung, J-Y., Lee, S-M., Shin, P-K., Chung, Y-C., 2000. Effect of pH on phase separated anaerobic digestion. Biotechnol. Bioprocess Eng. 5, 456–459. 24. Kardos, L., Juhasz, A., Palko, G.Y., Olah, J., Barkacs, K., Zaray, G., 2011. Comparing of mesophilic and thermophilic anaerobic fermented sewage sludge based on chemical and biochemical tests. Appl. Ecol. Environ. Res. 9, 293–302. 25. Latif, M.A., Metha, C.M., Batstone, D.J., 2017. Influence of low pH on continuous anaerobic digestion of waste activated sludge. Water Res. 113, 42– 49. 26. Lauwers, A.M., Heinen, W., Gorris, L.G.M., van der Drift, C., 1990. Early stage in biofilm development in methanogenic fluidized bed reactors. Appl. Microbiol. Biotechnol. 33, 352–358. 27. Lee, B., Park, J.G., Shin, W.B., Tian D.J., Jun, H.B., 2017. Microbial communities change in an anaerobic digestion after application of microbial electrolysis cells. Bioresour. Technol. 234, 273–280. 28. Logan, B.E., Call, D., Cheng, S., Hamelers, H.V., Sleutels, T.H., Jeremiasse, A.W., Rozendal, R.A., 2008. Microbial electrolysis cells for high yield hydrogen

22

gas production from organic matter. Environ. Sci. Technol. 42, 8630–8640. 29. Lopez, R.J., Higgins, S.R., Pagaling, E., Yan, T., Cooney, M.J., 2014. High rate anaerobic digestion of wastewater separated from grease trap waste. Renew. Energy. 62, 234–242. 30. Lu, J., Gavala, H.N., Skiadas, I.V., Mladenovska, Z., Ahring, B.K., 2008. Improving anaerobic sewage sludge digestion by implementation of a hyperthermophilic prehydrolysis step. J. Environ. Manage. 88, 881–889. 31. Ma, K., Liu, X., Dong, X., 2005. Methanobacterium beijingense sp. nov., a novel methanogen isolated from anaerobic digesters. Int. J. Syst. Evol. Microbiol. 55, 325–329. 32. Maus, I., Wibberg, D., Stantscheff, R., Cibis, K., Eikmeyer, F.G., König, H., Pühler, A., Schlüter, A., 2013. Complete genome sequence of the hydrogenotrophic Archaeon Methanobacterium sp. Mb1 isolated from a production-scale biogas plant. J. Biotechnol. 168, 734–736. 33. Mori, K., Harayama, S., 2011. Methanobacterium petrolearium sp. nov. and Methanobacterium ferruginis sp. nov., mesophilic methanogens isolated from salty environments. Int. J. Syst. Evol. Microbiol. 61, 138–143. 34. Moset, V., Bertolini, E., Cerisuelo, A., Cambra, M., Olmos, A., Cambra-López, M., 2014. Start-up strategies for thermophilic anaerobic digestion of pig manure. Energy. 74, 389–395. 35. Omidi, H., Sathasivan, A., 2013. Optimal temperature for microbes in an acetate fed microbial electrolysis cell (MEC). Int. Biodeterior. Biodegradation. 85, 688– 692. 36. Rajagopal, R., Massé, D.I., Singh, G., 2013. A critical review on inhibition of

23

anaerobic digestion process by excess ammonia. Bioresour. Technol. 143, 632– 641. 37. Siegert, I., Banks, C., 2005. The effect of volatile fatty acid additions on the anaerobic digestion of cellulose and glucose in bath reactors. Process Biochem. 40, 3412–3418. 38. Song, Y.C., Kim, D.S., Woo, J.H., 2014. Effect of epoxy mixed with Nafion solution as an anode binder on the performance of microbial fuel cell. J. Kor. Soc. Environ. Eng. 36, 1–6. 39. Tartakovsky, B., Metha, P., Bourque, J.S., Guiot, S.R., 2011. Electrolysisenhanced anaerobic digestion of wastewater. Bioresour. Technol. 102, 5685– 5691. 40. Van Bruggen, J.J.A., Zwart, K.B., van Assema, R.M., Stumm, C.K., Vogels, G.G., 1984. Methanobacterium formicicum, an endosymbiont of the anaerobic ciliate Metopus striatus McMurrich. Arch. Microbiol. 139, 1–7. 41. Wang, H., Zhang Y., Angelidaki, I., 2016. Ammonia inhibition on hydrogen enriched anaerobic digestion of manure under mesophilic and thermophilic conditions. Water Res. 105, 314–319. 42. Xafenias, N., Mapelli, V., 2014. Performance and bacterial enrichment of bioelectrochemical systems during methane and acetate production. Int. J. Hydrogen Energy. 39, 21864–21875. 43. Yi, H., Yong, D., Lee, K., Cho, Y. J., Chun, J., 2014. Profiling bacterial community in upper respiratory tracts. BMC Infect. Dis. 14, 583. 44. Yin, Q., Zhu, X., Zhan, G., Bo, T., Yang, Y., Tao, Y., He, X., Li, D., Yan, Z., 2016. Enhanced methane production in an anaerobic digestion and microbial

24

electrolysis cell coupled system with co-cultivation of Geobacter and Methanosarcina. J. Environ. Sci. 42, 210–214. 45. Zhang, J., Zhang, Y., Quan, X., Chen, S., Afzal, S., 2013. Enhanced anaerobic digestion of organic contaminants containing diverse microbial population by combined microbial electrolysis cell (MEC) and anaerobic reactor under Fe(III) reducing conditions. Bioresour. Technol. 136, 273–280. 46. Zhang, Y., Angelidaki, I., 2014. Microbial electrolysis cells turning to be versatile technology: Recent advances and future challenges. Water Res. 56, 11– 25. 47. Zhao, Z., Zhang, Y., Chen, S., Quan, X., Yu, Q., 2014. Bioelectrochemical enhancement of anaerobic methanogenesis for high organic load rate wastewater treatment in a up-flow anaerobic sludge blanket (UASB) reactor. Sci. Rep. 4, 6658. 48. Zinder, S.H., Sowers, K.R., Ferry, J.G., 1985. Methanosarcina thermophila sp. nov., a thermophilic, acetotrophic, methane-producing bacterium. Int. J. Syst. Bacteriol. 35, 522–523. 49. Zuo, Y., Xing, D., Regan, J.M., Logan, B.E., 2008. Isolation of the exoelectrogenic bacterium Ochrobactrum anthropic YZ-1 by using a U-tube microbial fuel cell. Appl. Environ. Microbiol. 74, 3130–3137.

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Figure Captions Figure 1. Configuration of (a) the two sequencing batch reactors and (b) electrodes (Nicoated graphite carbon anode and Cu-, Ni, and Fe-coated graphite carbon cathode). Figure 2. (a) Effluent TCOD and (b) daily methane production in the AD and AD + MEC reactors during the three experimental stages; start-up (1–69 d), intermediate steady state (70–289 d), and final steady state (290–365 d). Figure 3. Methane production rates of the AD and AD + MEC reactors during the three experimental stages ((a) start-up; 1–69 d, (b) intermediate steady state; 70–289 d, and (c) final steady state; 290–365 d). R2 values of AD and AD + MEC: (a) 0 and 0.9744, (b) 0.9527 and 0.9735, (c) 0.9999 and 0.9999. Figure 4. Methane yields based on TCOD removed during the three experimental stages. Average methane yield (L-CH4/g-CODrem) of the AD and AD + MEC: (a) 0 and 0.21 ± 0.09, (b) 0.28 ±0.05 and 0.33 ±0.02, (c) 0.33 ±0.02 and 0.34 ±0.02. Figure 5. Effects of the hydrogen ion (H+) concentration (a) and VFA concentration (b) on methane production in the AD and AD + MEC reactors. Figure 6. Microbial communities in the AD and AD + MEC reactors in the intermediate steady state (sludge sampled at 200 d). (a) Counts (number) and (b) proportions (%) of bacterial species in the reactors, (c) counts (number) and (d) proportions (%) of archaeal species in the reactors.

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Table 1. Operational conditions of the AD and AD + MEC reactors Reactors Parameters AD

AD + MEC

Inoculated sludge

AD reactor sludge from the FWTP

Voltage



Electrode material



0.3 V Anode: Ni-coated graphite carbon Cathode: Cu, Ni, and Fe-coated graphite carbon

Equipped electrodes



6 sets

OLR

3.0 kg-TCOD/m3d

HRT

20 days

Operation period Operation type Temperature

12 months Sequencing batch reactor 35C

AD, anaerobic digestion; MEC, microbial electrolysis cell; FWTP, food waste treatment plant; OLR, organic loading rate; TCOD, total chemical oxygen demand; HRT, hydraulic retention time

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Table 2. Properties of AD and AD + MEC reactors during the three experimental stages Parameters

AD reactor

AD + MEC reactor

Start-

Interm

Final

Start-

Interm

Final

up

ediate

steady

up

ediate

steady

steady

state

steady

state

state Methane

state 16.5

production 0.2 ±0.1 7.2 ±2.9

Biogas

CH4

Content (%) CO2

17.0

±2.2

±1.6

6.9 ±3.4 ±1.7

(L/d)

12.1

55.7

62.7

64.4

65.8

65.7

65.9

±7.5

±1.2

±1.3

±8.3

±1.4

±1.1

24.8

33.1

32.8

25.9

30.6

31.2

±4.8

±0.3

±1.1

±7.6

±1.7

±0.8

14.9

H2

3.3 ±1.9 2.3 ±0.8 4.1 ±1.8 2.8 ±0.5 2.4 ±0.6 ±2.9 TCOD

removal

eff.

(%) TVS removal eff. (%)

pH

NH4+-N (mg/L)

Alkalinity CaCO3)

(mg/L

as

28.1

59.0

74.7

64.9

71.9

76.1

±10.3

±6.5

±5.8

±4.2

±2.5

±3.3

32.9

61.6

72.7

59.7

69.6

73.2

±13.9

±3.1

±2.4

±8.0

±2.6

±2.1

6.67

7.62

7.86

7.42

7.67

7.93

±0.63

±0.22

±0.14

±0.18

±0.23

±0.11

2,216

1,721

1,467

1,570

1,434

1,398

±416

±115

±138

±83

±98

±141

4,333

6,510

7,866

6,885

7,197

8,050

±1,691

±234

±342

±774

±489

±327 28

VFAs

6,207 3,254±4

2,056

4,052

2,406

1,838

±772

69

±335

±534

±309

±173

210

173

62

108

53

38

±172

±52

±21

±59

±33

±19

496

222

118

243

67

99

±293

±224

±32

±40

±21

±30

633

149

96

452

73

98

acid

±250

±214

±74

±55

±32

±21

Butyric

4,894

2,710

1,752

3,246

2,172

1,623

acid

±772

±333

±292

±463

±128

±176

1.2

0.38

0.25

0.52

0.29

0.21

±0.79

±0.03

±0.04

±0.14

±0.03

±0.02

Total

concentra tion

Acetic

(mg/L)

acid Propionic acid Lactic

VFAs/Alk. ratio

AD, anaerobic digestion; MEC, microbial electrolysis cell; TCOD, total chemical oxygen demand; TVS, total volatile solid; VFAs, volatile fatty acids.

29

Figure 1

30

Figure 2

31

Figure 3

32

Figure 4

33

Figure 5

34

Figure 6

35

Highlights 

Steady-state methane production of AD and AD+MEC was near the theoretical maximum



MEC can accelerate the methane production rate and stabilization



MEC promotes the conversion of hydrogen ions and volatile fatty acids to methane



AD+MEC showed better organic matter removal via increased exoelectrogenic bacteria



MEC eliminates factors inhibiting acetoclastic methanogenesis in bulk

36