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Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment
Accepted Manuscript Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment Kai Cheng, Jingping Hu, Huijie Hou, Bingchuan Liu, Qin Chen, Keliang Pan, Wenhong Pu, Jiakuan Yang, Xu Wu, Changzhu Yang PII: DOI: Reference:
S0960-8524(17)30009-3 http://dx.doi.org/10.1016/j.biortech.2016.12.115 BITE 17491
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Bioresource Technology
Received Date: Revised Date: Accepted Date:
22 October 2016 30 December 2016 31 December 2016
Please cite this article as: Cheng, K., Hu, J., Hou, H., Liu, B., Chen, Q., Pan, K., Pu, W., Yang, J., Wu, X., Yang, C., Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment, Bioresource Technology (2017), doi: http://dx.doi.org/10.1016/j.biortech.2016.12.115
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Submitted to: Bioresource Technology
Date: 2016.12.30
Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment Kai Cheng#, Jingping Hu#, Huijie Hou, Bingchuan Liu, Qin Chen, Keliang Pan, Wenhong Pu, Jiakuan Yang, Xu Wu, Changzhu Yang * College of Environmental Science and Engineering, Huazhong University of Science and Technology, Luoyu Road 1037, Wuhan 430074, PR China
*Corresponding author: Prof. Changzhu Yang, Tel: +86-27-87792101, E-mail: [email protected]
#
These two authors contributed equally to the paper.
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Abstract Microbial consortiums aggregated on the anode surface of microbial fuel cells (MFCs) are critical factors for electricity generation as well as biodegradation efficiencies of organic compounds. Here in this study, aerobic granular sludge (AGS) was assembled on the surface of the MFC anode to form an AGS-MFC system with superior performance on epoxy reactive diluent (ERD) wastewater treatment. AGS-MFCs successfully shortened the startup time from 13 d to 7 d compared to the ones inoculated with domestic wastewater. Enhanced toxicity tolerance as well as higher COD removal (77.8% vs. 63.6%) were achieved. The higher ERD wastewater treatment efficiency of AGS-MFC is possibly attributed to the diverse microbial population on MFC biofilm, as well as the synergic degradation of contaminants by both the MFC anode biofilm and AGS granules. Keywords： Aerobic granular sludge; Microbial fuel cells; Recalcitrant organic pollutants; Biodegradation; Industrial wastewater
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1. Introduction The release of industrial wastewater containing recalcitrant organics could generate severe adverse impacts on both surface and ground water (Luo et al., 2009, Liu and Li, 2014, Abbasi et al., 2016, An et al., 2016, Wang et al., 2016), posing severe threat to both public health and ecological systems (Habibul et al., 2016). Current wastewater treatment plants for domestic wastewater treatment are mostly constructed based on the well-established activated sludge process. However, they are not suitable for the treatment of hazardous wastewater due to the susceptibility of microorganisms to the toxic components (Sahinkaya and Dilek, 2005). Microbial fuel cells have been intensively studied as novel platforms for both wastewater treatment and energy harvest (Liu et al., 2004; Ogugbue et al., 2015; Sonawane et al., 2014), demonstrating capabilities of combusting versatile organics from glucose (Chaudhuri and Lovley, 2003), acetate (Pant et al., 2010), sucrose (Kim et al., 2010) to domestic and brewery wastewater that contain more complicated mixed compounds (Kim et al., 2015; Wen et al., 2009). Furthermore, both organic and inorganic wastes originated from vegetable oil industries, glass and marble industries and chemical industries were also treated by MFCs, with 85%−90% of COD removals achieved (Abbasi et al., 2016). MFCs have also demonstrated their conceptual usages in recalcitrant contaminants removal using phenol as a typical treatment target (Pant et al., 2010). However, phenol degradation efficiency with closed circuit MFC was only 8-14% higher than that worked at opened circuit condition. 92.4% of the removed phenol was attributed to anaerobic digestion but not 3
exoelectrogenic microbial conversion. Thus, the advantages of using MFCs as a treatment strategy of phenol are very limited (Luo et al., 2009). In addition, a prolonged startup time for microbial biofilm acclimation is usually inevitable and low degradation efficiency as well as unstable performance are frequently observed. Degradation efficiency of a specific contaminant in an MFC is highly dependent on the microbial consortiums aggregated on the anode surface, which constitute electrochemical active bacteria that directly convert organic compounds into electricity, as well as non-electrochemical active bacteria that are also critical for electricity generation through synergetic cooperation. Source of inoculum and carbon source types are two critical parameters to determine the constitution of microbial consortiums of the electrogenic microbial biofilm. Metal reducing bacteria such as Geobacter spp., Shewanella spp. are typical microorganisms that are uncovered from MFC anode biofilm when natural river water or domestic wastewater is used as inoculum (Zhi et al., 2014, Hou et al., 2009). It is found that Desulfovibrio (18.4%) constitutes the key genus level content when brewery waste was used as the inoculum and azo dye as the carbon source (Miran et al., 2015). The microbial groups on anode biofilm were highly diverse, with 62.3% of all sequences unclassified. Feeding the MFC reactor with acetic acid decreased the diversity and richness of biofilm species, especially the richness of Geobacter sulfurreducens (Kiely et al., 2011). Diverse microbial communities in MFCs are desired to enhance electrogenic capability of biofilm for simultaneous carbon digestion and bioenergy generation (Kiely 4
et al., 2011), as well as to build up anti-toxic property enabling efficient reproduction of the microbial population. Previous study which used suspended sludge as microbial inoculum of MFCs was found to enhance power generation when treating feces wastewater (Zhi et al., 2008). Aerobic granular sludge (AGS) is considered to be a complicated system formed by large varieties of aerobic and anaerobic respiring microbes that colonized in different granule layers (Wen et al., 2009). They are densely-packed microbial aggregates with rich and strong microbial structure, high biomass retention, excellent biological efficiency and toxicity tolerance. Aerobic sludge granules have been successfully applied for the treatment of high strength organic and toxic industrial wastewater like phenol, 4-chlorophenol, etc. Aerobic granules played a promising role in adsorption of toxic chemicals and further degradation through co-metabolism of assorted microbial communities colonized in different layers of granules (Gao et al., 2011). The cultivation of aerobic granular sludge in MFC cathode chamber has been studied and the granule’s physiochemical properties were also investigated (Chen et al., 2014). However, strategies to enhance the efficiency of MFCs for recalcitrant wastewater treatment were very limited, and combining the advantages of both AGS and MFC technologies to simultaneously realize efficient recalcitrant wastewater treatment as well as enhancing bioenergy harvest has not been studied. Here in this study, the addition of AGS as the inoculum of MFCs was studied for the treatment of recalcitrant wastewater from epoxy reactive diluent (ERD) production. The startup time was evaluated compared to MFCs without AGS. Besides, toxicity tolerance 5
to the influent industrial wastewater was discussed. Both COD removal and power generation were obtained and compared to domestic wastewater inoculated MFCs. Mechanisms of the synergic effects of anodic biofilm and AGS on ERD wastewater treatment were also proposed.
2. Materials and methods 2.1 Aerobic granular sludge cultivation Mature aerobic granular sludge assembled in microbial fuel cells had been previously cultivated in a column sequencing batch reactor (SBR) reactor with a working volume of 9.9 L (diameter of 8.4 cm, height of 200 cm, H/D ratio of 21.4), a superficial gas velocity of 2.0 cm/s and a water exchange ratio of 61.8% at the end of each cycle (6 hours). 70% of activated sludge and 30% of previously cultivated aerobic granules were mixed and added into the SBR reactor to initiate the granulation. Synthesized wastewater consisted of sodium acetate, ammonia nitrogen, phosphates and minerals was fed. Other operation parameters were used as previously reported (Long et al., 2014).
2.2 MFC configuration and operation Carbon cloth anode (3.8 cm in diameter) was pretreated as follows: It was firstly cleaned by soaking in acetone, ethanol and deionized water sequentially with each for 15 min to remove possible organic contaminants. The cleaned carbon cloth was then heat-treated in a muffle furnace at 450 °C for 30 min. It was rinsed with deionized water three times 6
before used in an MFC (Feng et al., 2010). To fabricate the air-cathode (3 cm in diameter), 100 mg activated carbon powder,10 mg carbon black and 1 mL 10% (w/v) PVDF solution were mixed together, then the mixture was uniformly smeared on a stainless steel mesh and dipped in deionized water with the carbon side facing down for 15 min (Yang et al., 2014). Single-chamber air-cathode MFC reactor with a volume of 28 mL operated under batch mode as previously described was used (Liu and Logan, 2004). Aerobic granular sludge was weighed for 3 g and loaded on the surface of anode to form a layer covering the anode. Stainless steel mesh was gently pressed against the aerobic granular sludge to hold the AGS granules in place to prevent them from dropping off the anode surface while not breaking the granule structures (Fig. 1 and Fig. S1). MFC devices were fed with medium solution containing sodium acetate (1 g/L), 50 mL phosphate buffer solution (PBS), vitamins (5 mL/L) and minerals (12.5 mL/L) as previously described (Lovley and Phillips, 1988). During the start-up, AGS inoculated MFCs (AGS-MFCs) were fed with 28 mL medium solution, while MFCs without AGS were fed with 14 mL medium and 14 mL domestic municipal wastewater (DW-MFCs) from Tangxun Lake municipal wastewater treatment plant in Wuhan, China. Anode and cathode were connected with a 1000 Ω external resistor during the startup process. For ERD wastewater treatment, different proportions of ERD wastewater was mixed with 1 g/L sodium acetate synthetic wastewater to achieve ERD concentrations from 10% to 100% and added into both types of MFC sequentially. 7
The influent (sodium acetate) in startup period was changed whenever the voltage output dropped below 50 mV. In the steady power generation phase, sodium acetate substrate was exchanged every 2 days for both AGS-MFCs and DW-MFCs. For ERD wastewater treatment, a retention time of 36 hours was used for each cycle. Substrate was exchanged using a 10 mL syringe. All experiments were carried out in duplicates at room temperature (23 ± 1ºC).
2.3 Analytical method Voltages of MFCs were collected by an automatic data acquisition system (Keithley 2750, Tektronix, USA). Polarization curves were obtained by decreasing the external resistance from 30000 Ω to 50 Ω. Power density was obtained as previously described (Logan, et al., 2006). COD was measured according to the standard method (Eaton, A.D., et al. 1998. Standard methods for the examination of water and wastewater. Washington, DC, American Public Health Association). Columbic efficiency was calculated at a resistance of 1000 Ω based on changes of COD concentration in a batch cycle (Logan, et al., 2006). Variation of different replicates of each type of MFCs () was calculated by dividing the standard deviation () with the average of the maximum voltages of each repeats ( ) in the same cycle by = Std⁄ .
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3. Results and discussion 3.1 Start-up of AGS inoculated MFCs MFC startup time was dramatically decreased using AGS as the inoculum (AGS-MFCs). Voltage output reached 74 ± 20 mV at day 5, and a stable voltage generation of 393 ± 27 mV was achieved at day 8 (Fig. 2A). However, MFCs inoculated only with domestic wastewater showed no obvious electricity generation in the first 5 days and the voltage only reached 15 ± 6 mV at day 7. It took 13 days for the DW-MFC to reach the stable stage, producing a maximum voltage of 393 ± 5 mV. Previous studies showed that a lower concentration of phosphate buffer could induce faster anodic respiring bacteria colonization and alter the microbial consortium of the biofilm (Yanuka-Golub et al., 2016). Setting the anode at a more negative potential (−0.2 V) or using pre-acclimated culture is also preferable for cell acclimation (Zhang et al., 2013). However, adding additional amendments such as acetate, fumarate, glucose and Fe(III) are not beneficial to the enhancement of microbial biofilm acclimation as well as power generation (Liu et al., 2011). Different inoculums possessing varied microbial communities would directly affect the start-up of reactors (Rinaldi et al., 2008). Gao et al., (Gao et al., 2014) found that the aerobic sludge-seeded MFCs started up faster than the anaerobic sludge seeded MFCs, and the inoculation of aerobic sludge resulted in higher bacterial diversity and abundance in the anodic biofilm. AGS, holding layers of functional microbial communities that work synergistically during carbon source digestion, is similar to
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domestic wastewater in the way of providing abundant microbial sources for MFCs. What’s more, the microbial consortia were more functionally related along the carbon digestion chain, which would benefit the microbial growth on the anode to form exoelectrogenic biofilms. No obvious current generation was noticed at the very beginning stage in AGS-MFCs, revealing that the capability of electron transfer through direct contact from the AGS granules to the electrode was very limited and current generation of AGS-MFCs was mainly attributed to the rapid acclimated anode biofilm.
3.2 Electricity generation Similar voltage outputs were obtained by both types of MFCs at their stable stages, with 405 ± 10 mV and 404 ± 16 mV for AGS-MFCs and DW-MFCs respectively when sodium acetate was used as substrate (Fig. 2B). However, distinguished differences on voltage outputs were observed when ERD wastewater was added into both types of MFCs. No obvious voltage drop was noticed with 10% ERD addition for both AGS-MFCs and DW-MFCs, showing a considerable toxicity tolerance of both MFCs at lower ERD wastewater levels. However, sharp deterioration of voltage outputs were noticed in DW-MFCs with continuous increase of ERD concentrations in the influent (Fig. 2C). Voltage decreased from around 417 mV to 204 mV when ERD concentration was increased from 0% to 80% (a 51.1% drop), and the variance of the 3 repeats of the DW-MFCs also became severe (2.8 % vs. 13.6 %, with influent ERD of 10% and 80%
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respectively). For AGS-MFCs, a much better toxicity tolerance was noticed comparing to DW-MFCs (Fig. 2D). The voltage output was reasonably sustained compared to DW-MFCs, it dropped from 423 mV to 360 mV with ERD concentration increased from 0% to 80%, and the drop was lower than that of DW-MFCs (14.9% vs. 51.1% for AGS-MFCs and DW-MFCs respectively). The variance of the 3 replicates was also less significant than that of the DW-MFCs (2.0% vs. 13.6% for AGS-MFCs and DW-MFCs respectively) at 80% ERD. When 100% of ERD wastewater was injected, the voltage was slightly bounced back with DW-MFCs possibly due to microbial biofilm adapting to the influent chemicals. However, it’s still lower than that of the AGS-MFCs (299 mV for DW-MFCs vs. 341 mV for AGS-MFCs). ERD wastewater, which consists of epichlorohydrin and butylglycidylether, are known to generate biological toxicity towards microorganisms. The compact structure and biologically efficient aerobic granules with diverse microbial species are critical factors for its outstanding toxicity resistance. Aerobic granules covered on top of the biofilm electrode served as an effective shield for the exoelectrogenic biofilm from the attack of the toxic substrate, where the DW-MFC with its anode biofilm exposed unprotectively. This might be the main reason of less voltage degradation and better reproducibility of AGS-MFCs comparing to DW-MFCs. It is possible that if sufficient time was given to DW-MFC to adapt to the target recalcitrant contaminant, the microorganisms would acclimate to build up their toxicity tolerance. However, it usually take a unreasonably long acclimation time. Previous study showed that it took over 100 11
d in forming a microbial system that could degrade sulphamethoxazole to 3-amino-5-methylisoxazole due to the direct inhibition of the antibiotics on microbial activity (Wang et al., 2015), and still the anode biofilm may not be able to compete with the dense and highly functionally packed AGS granules when exposed to shock toxic contaminants. AGS-MFCs also demonstrated superior power generation capabilities when ERD was used as substrate compared to DW-MFCs. No obvious differences on maximum power densities when sodium acetate was used as the substrate, with 408 ± 26 mW/m2 and 404 ± 4 mW/m2 for AGS-MFCs and DW-MFCs respectively (Fig. 3A). Power densities of both types of MFCs were reduced when ERD was used as the substrate, with a 14 % decrease for AGS-MFCs and 29 % decrease for DW-MFCs, due to the toxicity of ERD compounds (Fig. 3A). AGS-MFCs showed max. power densities of 347 ±14 mW/m2, which was 21% higher than that of the DW-MFCs (287 ± 20 mW/m2). Since the electricity generation capabilities of both AGS-MFCs and DW-MFCs using sodium acetate as carbon source demonstrated comparable results, it is assumed that the abundance of exoelectrogens in the anode biofilm would be similar. This result is consistent with previous study using activated sludge and domestic wastewater as inoculums, where no power generation difference was observed when glucose was used as the carbon source (Zhi et al., 2008). Previous study showed that phenol was firstly biodegraded anaerobically to 4-hydroxybenzoic acid and 4-hydroxy-3-methylbenzoic acid, which were then converted to electricity by exoelectrogens accumulated on the 12
anode biofilm (Hedbavna et al., 2016). It indicated the cooperation among different microbial species besides exoelectrogens are necessary for efficient degradation of recalcitrant contaminants. More importantly, the hierarchical structure of AGS is an important factor of the AGS-MFC system being able to withstand the impact of toxins for stable power generation. Polarization curves showed that AGS-MFCs demonstrated slightly higher ohmic resistance than that of the DW-MFC when sodium acetate was used as inoculum (193 Ω with AGS vs. 188 Ω with DW) (Fig. 3B), showing that the AGS granules didn’t affect the internal ohmic resistance significantly. When changing the substrate from sodium acetate to ERD, AGS-MFCs kept the same resistance (193 Ω with sodium acetate vs. 194 Ω with ERD), however, DW-MFCs showed a substantial increase of the internal ohmic resistance (188 Ω with sodium acetate vs. 235 Ω with ERD), which might be another reason leading to the deterioration of the power generation in the DW-MFC systems. Power generation behaviors of both types of MFCs demonstrated that ERD wastewater addition showed obvious impact on bacterial activities, while the AGS-inoculated MFCs were appeared to be affected to a milder extent.
3.3 Wastewater treatment efficiency COD removal was greatly enhanced with AGS inoculated MFCs when ERD was used as the substrate. When sodium acetate was used as the carbon substrate, COD removals of AGS-MFCs were slightly higher than DW-MFCs at the early stage of MFC operation
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(80.8% and 70.5% at day 15), and they became identical as the operation time proceeded to 25 days and longer (82.1% for AGS-MFCs vs. 80.0% for DW-MFCs, 35 days) (Fig. 4A). As increased concentrations of ERD were used in the influent (from 10% to 80%), the effluent COD of DW-MFCs increased consistently (Fig. 4B), and the effluent COD of DW-MFC reached the maximum of 260.3 mg/L with a lowest COD removal of 51% when ERD was added in the proportion of 80%,. As for AGS-inoculated MFCs, effluent COD increased slightly but recovered in a short time. The highest effluent COD was 230 mg/L with a COD removal of 66%. With 100% ERD addition, the effluent remained to be relatively stable and COD was 118 mg/L with a removal rate of 77%. AGS-MFCs demonstrated better COD removal performance than that of the DW-MFCs. Considering the impact of toxic substance contained in the influent, MFCs inoculated with AGS greatly enhanced their tolerance to the wastewater toxicity. Similar trends were observed for coulombic efficiencies from day 15 to day 25, with AGS-MFCs showing slightly higher CE (26 ± 0.8 % vs. 24 ± 2 %) (Fig. 5). As the operation time proceeded, although identical COD removals were obtained at the same time by both MFC systems, AGS-MFC still showed minor advances over DW-MFCs on CE (28 ± 2 % vs. 26.9 ± 0.3 %).
3.4 Mechanisms of the synergic function of AGS and MFCs A stronger microbial aggregation is a valid defensive strategy against inhibiting effects towards toxic compounds (Jiang et al., 2002). The degradation process of pollutants,
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especially recalcitrant substances usually go through a long chain of decomposition, which is accomplished under the co-functions of multiple types of microbes. For aerobic granular sludge, layers formed at different distances away from granule core are colonized by different microbial species, which was mainly decided by oxygen and nutrients transportation. Anaerobic microorganisms in the central core and facultative microorganisms in the middle layer are not able to live in a substantial number in domestic water and suspended activated sludge. Huge quantity of microbes and rich biodiversity contained in AGS facilitate biofilm formation and enable rapid MFC startup (Fig. S2). Filamentous bacteria interacted with inorganic substance formed the out layer. Bacillius, Grevibacteria formed a microbial net in the middle layer. AGS-attached anode could protect electrogenic microbe from excessive exposures when AGS adsorbed toxins and degraded them into small or less-toxic substrates. In the core of AGS, irregular-shaped microbes could degrade substrate coming from the channel pores in each layer. Bacillus, Brevibacteria, filamentous bacteria were connected in a network by pili, which could serve as nanowire to transfer electrons. Meanwhile, Coccobacteria are tightly attached to the surface of large pylori. The rich and biodiverse microbial community structure is conducive to the operation of MFCs (Du et al., 2007), which is also proved by the minor internal ohmic resistance change by covering the anode electrode. Organic pollutants can be grouped into two categories according to the biotransformation process of microbes: the first category is growth substrate, which 15
provides energy and carbon for cell growth directly when degraded in metabolism. The second category is non-growth substrate, which contributes little or negligible amount of energy or carbon for cell growth. They are generally some pollutants from industrial effluent, hazardous wastes and etc. Non-growth substrate could be degraded in the presence of growth substrate, in which growth substrate serves as both energy provider and electron donor (Bali and Şengül, 2002). Co-substrate is known to affect the simultaneous removal of the target contaminant pentachlorophenol (PCP) in MFCs (Huang et al., 2011). Simultaneous addition of glucose could enhance the system tolerance of PCP shock as well as decrease the acclimation time (Huang et al., 2011). Sodium acetate is one of the most favorable target of microorganisms in wastewater (Kalishwaralal et al., 2009). When the mixture of ERD wastewater and synthesized wastewater (sodium acetate) was used as the influent, sodium acetate facilitated the wastewater treatment by serving as carbon source of microorganisms as well as co-metabolism reagent between electrogenic biofilm and AGS granules. Here in this research, adding sodium acetate as the co-substrate for ERD wastewater treatment didn’t affect the effluent COD in a dramatic way for AGS-MFCs, while a slight overall increase of effluent COD was noticed for DW-MFCs with less sodium acetate (lower effluent COD with 10%-50% ERD vs. higher effluent COD with 50%-100% ERD, Fig. 4B), showing that a higher sodium acetate concentration led to a better final water quality in terms of COD when domestic wastewater was used as the inoculum. Similar power generation, COD removal and coulombic efficiency were obtained for both types 16
of MFCs when only sodium acetate was used as substrate, revealing that biofilm itself was sufficient enough to degrade sodium acetate, a simpler form of carbon source. However, the superior ERD wastewater treatment performance achieved by the AGS-MFCs over DW-MFCs is different from that using sodium acetate as the model substrate (similar results achieved by both systems). This suggests that sodium acetate is not an optimized substrate for MFC performance evaluation to predict the system efficiency when treating more complex wastewaters. Finally, in the AGS-MFC hybrid system, AGS and MFCs could function synergistically to achieve a better performance on both wastewater treatment and bioenergy harvest (Scheme 1). First, AGS with its good adsorption ability could protect anodic biofilm from excessive exposure to toxins in influent (Scheme 1, mechanism ①). AGS is promising for toxic chemical adsorption due to its intrinsic characteristics as high surface area, porosity, hydrophobic property, and rich extracellular polymeric substances. Aerobic granular sludge was prepared as adsorbent to remove heavy metals (Cu 2+) (Adav et al., 2008, Jian et al., 2014). In sewage treatment, the removal of some pharmaceuticals (e.g., diazepam, diclofenac, ibuprofen, naproxen, sulfamethoxazole) was also mainly due to the adsorption of those compounds to sludge. Second, AGS, which serves as a bio-adsorbent in front of biofilm could preliminarily degrade organic compounds into smaller forms of compounds that can be easily decomposed (Scheme 1, mechanism ②). In wastewater treatment, the accessibility of pollutants to microorganisms can be defined in terms of external and internal bioavailability. In this process, physicochemical aspects 17
related to phase distribution and mass transfer, and physiological aspects related to microorganisms such as the presence of active uptake systems, their enzymatic equipment and ability to excrete enzymes, work unitedly on the degradation of pollutants. Microorganisms can transform organic molecules via the succession of oxidation reactions to simpler products (Meer, 2006). Electro-active bacteria attached on the anode then produce electricity through syntrophic interactions through carbon flow (Scheme 1, mechanism③). Further studies of microbial population as well as biomass of the biofilm in a simpler system (e.g. with a single type of carbon source) would be helpful for a better understanding of the synergic effect between AGS and exoelectrogens on recalcitrant organics degradation.
4. Conclusion The addition of AGS covering MFC anode provided for rapid reactor startup as well as efficient treatment of epoxy reactive diluent wastewater. Maximum power density of 491±8 mW/m2 and COD removal of 82±5% were achieved. AGS-MFCs didn't show dramatic advantages on sodium acetate combustion compared to DW-MFCs, however, superior advantages on ERD wastewater treatment were demonstrated due to the combination of AGS’s high toxicity tolerance, its optimized microbial population on biofilm and MFC anode providing readily available electron acceptor. The novel study provided a feasible method for industrial wastewater treatment combining the
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advantages of AGS and MFCs.
Acknowledgements This work was supported by the National Thousand Talents Program, Natural Science Foundation of China (51508213, 51608217 and 21607046), key project of Hubei Provincial Natural Science Foundation (2014CFA109), General program of Natural Science Foundation of Hubei Province (2016CFB539), Innovative and Interdisciplinary Team at HUST (0118261077) and Independent Innovation Foundation of HUST-Exploration Fund (2016YXMS288). The authors would like to thank the Analytical and Testing Center of Huazhong University of Science and Technology for providing the facilities and conduct the characterization work.
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References 1.
Abbasi, U., Jin, W., Pervez, A., Bhatti, Z.A., Tariq, M., Shaheen, S., Iqbal, A., Mahmood, Q., 2016. Anaerobic microbial fuel cell treating combined industrial wastewater: Correlation of electricity generation with pollutants. Bioresour. Technol., 200, 1-7.
2.
Adav, S.S., Lee, D.J., Show, K.Y., Tay, J.H., 2008. Aerobic granular sludge: Recent advances. Biotechnol. Adv., 26, 411-23.
3.
An, B.M., Heo, Y., Maitlo, H.A., Park, J.Y., 2016. Scaled-up dual anode/cathode microbial fuel cell stack for actual ethanolamine wastewater treatment. Bioresour. Technol., 210, 68-73.
4.
Bali, U., Şengül, F., 2002. Performance of a fed-batch reactor treating a wastewater containing 4-chlorophenol. Process Biochem., 37, 1317-1323.
5.
Chaudhuri, S.K., Lovley, D.R., 2003. Electricity generation by direct oxidation of glucose in mediatorless microbial fuel cells. Nature Biotechnol., 21, 1229-32.
6.
Chen T. Aerobic granular sludge stabilization in biocathode chamber of newly constructed continue flow microbial fuel cell system treating synthetic and pharmaceutical wastewater. Desali. Water Treat., 2014:1-10.
7.
Du, Z., Li, H., Gu, T., 2007. A state of the art review on microbial fuel cells: A promising technology for wastewater treatment and bioenergy. Biotechnol. Adv., 25,
20
464-82. 8.
Eaton, A.D., Lenore S.C., Arnold E.G., and Mary Ann H.F., 1998. Standard methods for the examination of water and wastewater. Washington, DC, American Public Health Association.
9.
Feng, Y., Yang, Q., Wang, X., Logan, B.E., 2010. Treatment of carbon fiber brush anodes for improving power generation in air–cathode microbial fuel cells. J Power Sources, 195, 1841-1844.
10. Gao, C.Y., Wang, A.J., Wu, W.M., Yin, Y.L., Zhao, Y.G., 2014. Enrichment of anodic biofilm inoculated with anaerobic or aerobic sludge in single chambered air-cathode microbial fuel cells. Bioresour. Technol., 167C, 124-132. 11. Gao, D., Liu, L., Liang, H., Wu, W., 2011. Aerobic granular sludge: characterization, mechanism of granulation and application to wastewater treatment. Crit. Rev. Biotechnol., 31, 137-152. 12. Habibul, N., Hu, Y., Wang, Y., Chen, W., Yu, H., Sheng, G., 2016. Bioelectrochemical Chromium (VI) Removal in Plant-Microbial Fuel Cells. Environ. Sci. Technol., 50, 3882-3889. 13. Hou, H., Li, L., Cho, Y., De Figueiredo, P., Han, A., 2009. Microfabricated microbial fuel cell arrays reveal electrochemically active microbes. PLOS one, 4, e6570. 14. Huang, L., Gan, L., Zhao, Q., Logan, B.E., Lu, H., Chen, G., 2011. Degradation of pentachlorophenol with the presence of fermentable and non-fermentable 21
co-substrates in a microbial fuel cell. Bioresour. Technol., 102, 8762-8. 15. Jian M, Tang C, Liu M. Adsorptive removal of Cu2+ from aqueous solution using aerobic granular sludge. Desali. Water Treat., 2014, 54(7):1-10. 16. Jiang, H.L., Tay, J.H., Tay, T.L., 2002. Aggregation of immobilized activated sludge cells into aerobically grown microbial granules for the aerobic biodegradation of phenol. Lett. Appl. Microb. 35, 439-45. 17. Kalishwaralal, K., Deepak, V., Pandian, S.R.K., Gurunathan, S., 2009. Biological synthesis of gold nanocubes from Bacillus licheniformis. Bioresour. Technol., 100, 5356-5358. 18. Kiely, P.D., Cusick, R., Call, D.F., Selembo, P.A., Regan, J.M., Logan, B.E., 2011. Anode microbial communities produced by changing from microbial fuel cell to microbial electrolysis cell operation using two different wastewaters. Bioresour. Technol., 102, 388-394. 19. Kim, J.R., Premier, G.C., Hawkes, F.R., Rodríguez, J., Dinsdale, R.M., Guwy, A.J., 2010. Modular tubular microbial fuel cells for energy recovery during sucrose wastewater treatment at low organic loading rate. Bioresour. Technol., 101, 1190-1198. 20. Kim, K.Y., Yang, W., Logan, B.E., 2015. Impact of electrode configurations on retention time and domestic wastewater treatment efficiency using microbial fuel cells. Water Res., 80, 41-46. 22
21. Liu, B., Li, B., 2014. Single chamber microbial fuel cells (SCMFCs) treating wastewater containing methanol. Int. J Hydrogen Energ., 39, 2340-2344. 22. Liu, G., Yates, M.D., Cheng, S., Call, D.F., Sun, D., Logan, B.E., 2011. Examination of microbial fuel cell start-up times with domestic wastewater and additional amendments. Bioresour. Technol., 102, 7301-6. 23. Liu, H., Logan, B.E., 2004. Electricity generation using an air-cathode single chamber microbial fuel cell in the presence and absence of a proton exchange membrane. Environ. Sci. Technol., 38, 4040-6. 24. Liu, H., Ramnarayanan, R., Logan, B.E., 2004. Production of electricity during wastewater treatment using a single chamber microbial fuel cell. Environ. Sci. Technol., 38, 2281-2285. 25. Logan, B.E., Hamelers, B., Rozendal, R., SchrÖder, U., Keller, J., Freguia, S, Aelterman, P., Verstraete, W., Rabaey, K., 2006. Microbial fuel cells: Methodology and Technology.
Environ. Sci. Technol., 40, 5181-5192
26. Long, B., Yang, C., Pu, W., Yang, J., Jiang, G., Dan, J., Li, C., Liu, F., 2014. Rapid cultivation of aerobic granular sludge in a pilot scale sequencing batch reactor. Bioresour. Technol., 166, 57-63. 27. Lovley, D.R., Phillips, E.J., 1988. Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese. Appl. Environ. Microb., 54, 1472-1480. 23
28. Luo, H., Liu, G., Zhang, R., Jin, S., 2009. Phenol degradation in microbial fuel cells. Chem. Eng. J., 147, 259-264. 29. Meer, J.R.V.D., 2006. Environmental pollution promotes selection of microbial degradation pathways. Front. Ecol. Environ., 4, 35-42. 30. Miran, W., Nawaz, M., Kadam, A., Shin, S., Heo, J., Jang, J., Lee, D.S., 2015. Microbial community structure in a dual chamber microbial fuel cell fed with brewery waste for azo dye degradation and electricity generation. Environ. Sci. Pollut. R., 22, 13477-13485. 31. Ogugbue, C., Ebode, E., Leera, S., 2015. Electricity generation from swine wastewater using microbial fuel cell., Water Res., 16, 26-33. 32. Pant, D., Bogaert, G.V., Smet, M.D., Diels, L., Vanbroekhoven, K., 2010. Use of novel permeable membrane and air cathodes in acetate microbial fuel cells. Electrochim. Acta, 55, 7710-7716. 33. Pant, D., Van Bogaert, G., Diels, L., Vanbroekhoven, K., 2010. A review of the substrates used in microbial fuel cells (MFCs) for sustainable energy production. Bioresour. Technol., 101, 1533-1543. 34. Rinaldi, A., Mecheri, B., Garavaglia, V., Licoccia, S., Di Nardo, P., Traversa, E., 2008. Engineering materials and biology to boost performance of microbial fuel cells: a critical review. Energ. Environ. Sci., 1, 417.
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35. Sahinkaya, E., Dilek, F.B., 2005. Biodegradation of 4-chlorophenol by acclimated and unacclimated activated sludge—Evaluation of biokinetic coefficients. Environ. Res., 99, 243-252. 36. Sonawane, J.M., Marsili, E., Ghosh, P.C., 2014. Treatment of domestic and distillery wastewater in high surface microbial fuel cells. Int. J. Hydrogen Energ., 39, 21819-21827. 37. Wang, C., Deng, H., Zhao, F., 2016. The remediation of chromium (VI)-contaminated soils using microbial fuel cells. Soil Sediment Contaminat., 25, 1-12. 38. Wen, Q., Ying, W., Cao, D., Zhao, L., Qian, S., 2009. Electricity generation and modeling of microbial fuel cell from continuous beer brewery wastewater. Bioresour. Technol., 100, 4171-4175. 39. Yang, W., He, W., Zhang, F., Hickner, M.A., Logan, B.E., 2014. Single-Step Fabrication Using a Phase Inversion Method of Poly (vinylidene fluoride) (PVDF) Activated Carbon Air Cathodes for Microbial Fuel Cells. Environ. Sci. Technol. Lett., 1, 416-420. 40. Yanuka-Golub, K., Reshef, L., Rishpon, J., Gophna, U., 2016. Community structure dynamics during startup in microbial fuel cells – the effect of phosphate concentrations. Bioresour. Technol., 212, 151-159. 41. Zhang, F., Xia, X., Luo, Y., Sun, D., Call, D.F., Logan, B.E., 2013. Improving startup 25
performance with carbon mesh anodes in separator electrode assembly microbial fuel cells. Bioresour. Technol., 133, 74-81. 42. Zhi, W., Ge, Z., He, Z., Zhang, H., 2014. Methods for understanding microbial community structures and functions in microbial fuel cells: a review. Bioresour. Technol., 171, 461-468. 43. Zhi, Y., Liu, H., Yao, L., 2008. The Effect of Suspended Sludge on Electricity Generation in Microbial Fuel Cells. Int. Conf. Bioinform. Biomed. Eng., pp. 2923-2927
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Figure Captions Fig. 1 Setup of the AGS-MFC system: a dense layer of AGS covering the MFC anode electrode. The AGS layer is held in place with a stainless steel mesh;
Fig. 2 (A) Startup of both DW-MFCs and AGS-MFCs; (B) Stable power generation of AGS-MFCs using 1 g/L sodium acetate as carbon source; (C) Voltage generation of DW-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate; (D) Voltage generation of AGS-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate.
Fig. 3 (A) Power curves and (B) polarization curves of DW/AGS-MFCs treating different types of wastewater (Sodium acetate-NaAc and ERD wastewater-ERD).
Fig. 4 COD concentrations in the influent and the effluent and COD removal of DW-MFCs and AGS-MFCs treating different types of wastewater: (A) synthesized waster (1 g/L sodium acetate) and (B) ERD wastewater with an increasing proportion.
Fig. 5 Coulombic efficiencies of DW-MFCs and AGS-MFCs treating different types of wastewater. (A) ERD wastewater and (B) synthesized waster (1 g/L sodium acetate).
Scheme 1 Proposed mechanisms of rapid AGS-MFC startup by facilitating biofilm formation as well as synergetic biodegradation of toxins by MFC biofilm anode with AGS granules during ERD wastewater treatment (①: AGS adsorbs a certain amount of
27
toxins in the solution, forming a protective layer to prohibit the direct attack of toxins towards the biofilm anode; ②: AGS degrades toxins into more biodegradable components that could penetrate the AGS layer to the biofilm surface; ③: Bioelectricity generation by electrogenic biofilm anode converting carbon source to electrons).
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Figures
Fig. 1 Setup of the AGS-MFC system: a dense layer of AGS covering the MFC anode electrode. The AGS layer is held in place with a stainless steel mesh;
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Fig. 2 (A) Startup of both DW-MFCs and AGS-MFCs; (B) Stable power generation of AGS-MFCs using 1 g/L sodium acetate as carbon source; (C) Voltage generation of DW-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate; (D) Voltage generation of AGS-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate.
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Fig. 4 COD concentrations in the influent and the effluent and COD removal of DW-MFCs and AGS-MFCs treating different types of wastewater: (A) synthesized waster (1 g/L sodium acetate) and (B) ERD wastewater with an increasing proportion.
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Fig. 5 Coulombic efficiencies of DW-MFCs and AGS-MFCs treating different types of wastewater. (A) ERD wastewater and (B) synthesized waster (1 g/L sodium acetate).
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Scheme 1 Proposed mechanisms of rapid AGS-MFC startup by facilitating biofilm formation as well as synergetic biodegradation of toxins by MFC biofilm anode with AGS granules during ERD wastewater treatment (①: AGS adsorbs a certain amount of toxins in the solution, forming a protective layer to prohibit the direct attack of toxins towards the biofilm anode; ②: AGS degrades toxins into more biodegradable components that could penetrate the AGS layer to the biofilm surface; ③: Bioelectricity generation by electrogenic biofilm anode converting carbon source to electrons).
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Graphical Abstract
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Highlights •
Aerobic granular sludge was incorporated into MFCs to form a hybrid system.
•
AGS-MFCs shortened the startup time from 13 d to 7 d compared to MFCs without
AGS. •
Superior performance for epoxy reactive diluent wastewater treatment was achieved.
•
The high efficiency is possibly due to the synergic effect of both MFC and AGS.
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S0960-8524(17)30009-3 http://dx.doi.org/10.1016/j.biortech.2016.12.115 BITE 17491
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
22 October 2016 30 December 2016 31 December 2016
Please cite this article as: Cheng, K., Hu, J., Hou, H., Liu, B., Chen, Q., Pan, K., Pu, W., Yang, J., Wu, X., Yang, C., Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment, Bioresource Technology (2017), doi: http://dx.doi.org/10.1016/j.biortech.2016.12.115
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
Submitted to: Bioresource Technology
Date: 2016.12.30
Aerobic granular sludge inoculated microbial fuel cells for enhanced epoxy reactive diluent wastewater treatment Kai Cheng#, Jingping Hu#, Huijie Hou, Bingchuan Liu, Qin Chen, Keliang Pan, Wenhong Pu, Jiakuan Yang, Xu Wu, Changzhu Yang * College of Environmental Science and Engineering, Huazhong University of Science and Technology, Luoyu Road 1037, Wuhan 430074, PR China
*Corresponding author: Prof. Changzhu Yang, Tel: +86-27-87792101, E-mail: [email protected]
#
These two authors contributed equally to the paper.
1
Abstract Microbial consortiums aggregated on the anode surface of microbial fuel cells (MFCs) are critical factors for electricity generation as well as biodegradation efficiencies of organic compounds. Here in this study, aerobic granular sludge (AGS) was assembled on the surface of the MFC anode to form an AGS-MFC system with superior performance on epoxy reactive diluent (ERD) wastewater treatment. AGS-MFCs successfully shortened the startup time from 13 d to 7 d compared to the ones inoculated with domestic wastewater. Enhanced toxicity tolerance as well as higher COD removal (77.8% vs. 63.6%) were achieved. The higher ERD wastewater treatment efficiency of AGS-MFC is possibly attributed to the diverse microbial population on MFC biofilm, as well as the synergic degradation of contaminants by both the MFC anode biofilm and AGS granules. Keywords： Aerobic granular sludge; Microbial fuel cells; Recalcitrant organic pollutants; Biodegradation; Industrial wastewater
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1. Introduction The release of industrial wastewater containing recalcitrant organics could generate severe adverse impacts on both surface and ground water (Luo et al., 2009, Liu and Li, 2014, Abbasi et al., 2016, An et al., 2016, Wang et al., 2016), posing severe threat to both public health and ecological systems (Habibul et al., 2016). Current wastewater treatment plants for domestic wastewater treatment are mostly constructed based on the well-established activated sludge process. However, they are not suitable for the treatment of hazardous wastewater due to the susceptibility of microorganisms to the toxic components (Sahinkaya and Dilek, 2005). Microbial fuel cells have been intensively studied as novel platforms for both wastewater treatment and energy harvest (Liu et al., 2004; Ogugbue et al., 2015; Sonawane et al., 2014), demonstrating capabilities of combusting versatile organics from glucose (Chaudhuri and Lovley, 2003), acetate (Pant et al., 2010), sucrose (Kim et al., 2010) to domestic and brewery wastewater that contain more complicated mixed compounds (Kim et al., 2015; Wen et al., 2009). Furthermore, both organic and inorganic wastes originated from vegetable oil industries, glass and marble industries and chemical industries were also treated by MFCs, with 85%−90% of COD removals achieved (Abbasi et al., 2016). MFCs have also demonstrated their conceptual usages in recalcitrant contaminants removal using phenol as a typical treatment target (Pant et al., 2010). However, phenol degradation efficiency with closed circuit MFC was only 8-14% higher than that worked at opened circuit condition. 92.4% of the removed phenol was attributed to anaerobic digestion but not 3
exoelectrogenic microbial conversion. Thus, the advantages of using MFCs as a treatment strategy of phenol are very limited (Luo et al., 2009). In addition, a prolonged startup time for microbial biofilm acclimation is usually inevitable and low degradation efficiency as well as unstable performance are frequently observed. Degradation efficiency of a specific contaminant in an MFC is highly dependent on the microbial consortiums aggregated on the anode surface, which constitute electrochemical active bacteria that directly convert organic compounds into electricity, as well as non-electrochemical active bacteria that are also critical for electricity generation through synergetic cooperation. Source of inoculum and carbon source types are two critical parameters to determine the constitution of microbial consortiums of the electrogenic microbial biofilm. Metal reducing bacteria such as Geobacter spp., Shewanella spp. are typical microorganisms that are uncovered from MFC anode biofilm when natural river water or domestic wastewater is used as inoculum (Zhi et al., 2014, Hou et al., 2009). It is found that Desulfovibrio (18.4%) constitutes the key genus level content when brewery waste was used as the inoculum and azo dye as the carbon source (Miran et al., 2015). The microbial groups on anode biofilm were highly diverse, with 62.3% of all sequences unclassified. Feeding the MFC reactor with acetic acid decreased the diversity and richness of biofilm species, especially the richness of Geobacter sulfurreducens (Kiely et al., 2011). Diverse microbial communities in MFCs are desired to enhance electrogenic capability of biofilm for simultaneous carbon digestion and bioenergy generation (Kiely 4
et al., 2011), as well as to build up anti-toxic property enabling efficient reproduction of the microbial population. Previous study which used suspended sludge as microbial inoculum of MFCs was found to enhance power generation when treating feces wastewater (Zhi et al., 2008). Aerobic granular sludge (AGS) is considered to be a complicated system formed by large varieties of aerobic and anaerobic respiring microbes that colonized in different granule layers (Wen et al., 2009). They are densely-packed microbial aggregates with rich and strong microbial structure, high biomass retention, excellent biological efficiency and toxicity tolerance. Aerobic sludge granules have been successfully applied for the treatment of high strength organic and toxic industrial wastewater like phenol, 4-chlorophenol, etc. Aerobic granules played a promising role in adsorption of toxic chemicals and further degradation through co-metabolism of assorted microbial communities colonized in different layers of granules (Gao et al., 2011). The cultivation of aerobic granular sludge in MFC cathode chamber has been studied and the granule’s physiochemical properties were also investigated (Chen et al., 2014). However, strategies to enhance the efficiency of MFCs for recalcitrant wastewater treatment were very limited, and combining the advantages of both AGS and MFC technologies to simultaneously realize efficient recalcitrant wastewater treatment as well as enhancing bioenergy harvest has not been studied. Here in this study, the addition of AGS as the inoculum of MFCs was studied for the treatment of recalcitrant wastewater from epoxy reactive diluent (ERD) production. The startup time was evaluated compared to MFCs without AGS. Besides, toxicity tolerance 5
to the influent industrial wastewater was discussed. Both COD removal and power generation were obtained and compared to domestic wastewater inoculated MFCs. Mechanisms of the synergic effects of anodic biofilm and AGS on ERD wastewater treatment were also proposed.
2. Materials and methods 2.1 Aerobic granular sludge cultivation Mature aerobic granular sludge assembled in microbial fuel cells had been previously cultivated in a column sequencing batch reactor (SBR) reactor with a working volume of 9.9 L (diameter of 8.4 cm, height of 200 cm, H/D ratio of 21.4), a superficial gas velocity of 2.0 cm/s and a water exchange ratio of 61.8% at the end of each cycle (6 hours). 70% of activated sludge and 30% of previously cultivated aerobic granules were mixed and added into the SBR reactor to initiate the granulation. Synthesized wastewater consisted of sodium acetate, ammonia nitrogen, phosphates and minerals was fed. Other operation parameters were used as previously reported (Long et al., 2014).
2.2 MFC configuration and operation Carbon cloth anode (3.8 cm in diameter) was pretreated as follows: It was firstly cleaned by soaking in acetone, ethanol and deionized water sequentially with each for 15 min to remove possible organic contaminants. The cleaned carbon cloth was then heat-treated in a muffle furnace at 450 °C for 30 min. It was rinsed with deionized water three times 6
before used in an MFC (Feng et al., 2010). To fabricate the air-cathode (3 cm in diameter), 100 mg activated carbon powder,10 mg carbon black and 1 mL 10% (w/v) PVDF solution were mixed together, then the mixture was uniformly smeared on a stainless steel mesh and dipped in deionized water with the carbon side facing down for 15 min (Yang et al., 2014). Single-chamber air-cathode MFC reactor with a volume of 28 mL operated under batch mode as previously described was used (Liu and Logan, 2004). Aerobic granular sludge was weighed for 3 g and loaded on the surface of anode to form a layer covering the anode. Stainless steel mesh was gently pressed against the aerobic granular sludge to hold the AGS granules in place to prevent them from dropping off the anode surface while not breaking the granule structures (Fig. 1 and Fig. S1). MFC devices were fed with medium solution containing sodium acetate (1 g/L), 50 mL phosphate buffer solution (PBS), vitamins (5 mL/L) and minerals (12.5 mL/L) as previously described (Lovley and Phillips, 1988). During the start-up, AGS inoculated MFCs (AGS-MFCs) were fed with 28 mL medium solution, while MFCs without AGS were fed with 14 mL medium and 14 mL domestic municipal wastewater (DW-MFCs) from Tangxun Lake municipal wastewater treatment plant in Wuhan, China. Anode and cathode were connected with a 1000 Ω external resistor during the startup process. For ERD wastewater treatment, different proportions of ERD wastewater was mixed with 1 g/L sodium acetate synthetic wastewater to achieve ERD concentrations from 10% to 100% and added into both types of MFC sequentially. 7
The influent (sodium acetate) in startup period was changed whenever the voltage output dropped below 50 mV. In the steady power generation phase, sodium acetate substrate was exchanged every 2 days for both AGS-MFCs and DW-MFCs. For ERD wastewater treatment, a retention time of 36 hours was used for each cycle. Substrate was exchanged using a 10 mL syringe. All experiments were carried out in duplicates at room temperature (23 ± 1ºC).
2.3 Analytical method Voltages of MFCs were collected by an automatic data acquisition system (Keithley 2750, Tektronix, USA). Polarization curves were obtained by decreasing the external resistance from 30000 Ω to 50 Ω. Power density was obtained as previously described (Logan, et al., 2006). COD was measured according to the standard method (Eaton, A.D., et al. 1998. Standard methods for the examination of water and wastewater. Washington, DC, American Public Health Association). Columbic efficiency was calculated at a resistance of 1000 Ω based on changes of COD concentration in a batch cycle (Logan, et al., 2006). Variation of different replicates of each type of MFCs () was calculated by dividing the standard deviation () with the average of the maximum voltages of each repeats ( ) in the same cycle by = Std⁄ .
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3. Results and discussion 3.1 Start-up of AGS inoculated MFCs MFC startup time was dramatically decreased using AGS as the inoculum (AGS-MFCs). Voltage output reached 74 ± 20 mV at day 5, and a stable voltage generation of 393 ± 27 mV was achieved at day 8 (Fig. 2A). However, MFCs inoculated only with domestic wastewater showed no obvious electricity generation in the first 5 days and the voltage only reached 15 ± 6 mV at day 7. It took 13 days for the DW-MFC to reach the stable stage, producing a maximum voltage of 393 ± 5 mV. Previous studies showed that a lower concentration of phosphate buffer could induce faster anodic respiring bacteria colonization and alter the microbial consortium of the biofilm (Yanuka-Golub et al., 2016). Setting the anode at a more negative potential (−0.2 V) or using pre-acclimated culture is also preferable for cell acclimation (Zhang et al., 2013). However, adding additional amendments such as acetate, fumarate, glucose and Fe(III) are not beneficial to the enhancement of microbial biofilm acclimation as well as power generation (Liu et al., 2011). Different inoculums possessing varied microbial communities would directly affect the start-up of reactors (Rinaldi et al., 2008). Gao et al., (Gao et al., 2014) found that the aerobic sludge-seeded MFCs started up faster than the anaerobic sludge seeded MFCs, and the inoculation of aerobic sludge resulted in higher bacterial diversity and abundance in the anodic biofilm. AGS, holding layers of functional microbial communities that work synergistically during carbon source digestion, is similar to
9
domestic wastewater in the way of providing abundant microbial sources for MFCs. What’s more, the microbial consortia were more functionally related along the carbon digestion chain, which would benefit the microbial growth on the anode to form exoelectrogenic biofilms. No obvious current generation was noticed at the very beginning stage in AGS-MFCs, revealing that the capability of electron transfer through direct contact from the AGS granules to the electrode was very limited and current generation of AGS-MFCs was mainly attributed to the rapid acclimated anode biofilm.
3.2 Electricity generation Similar voltage outputs were obtained by both types of MFCs at their stable stages, with 405 ± 10 mV and 404 ± 16 mV for AGS-MFCs and DW-MFCs respectively when sodium acetate was used as substrate (Fig. 2B). However, distinguished differences on voltage outputs were observed when ERD wastewater was added into both types of MFCs. No obvious voltage drop was noticed with 10% ERD addition for both AGS-MFCs and DW-MFCs, showing a considerable toxicity tolerance of both MFCs at lower ERD wastewater levels. However, sharp deterioration of voltage outputs were noticed in DW-MFCs with continuous increase of ERD concentrations in the influent (Fig. 2C). Voltage decreased from around 417 mV to 204 mV when ERD concentration was increased from 0% to 80% (a 51.1% drop), and the variance of the 3 repeats of the DW-MFCs also became severe (2.8 % vs. 13.6 %, with influent ERD of 10% and 80%
10
respectively). For AGS-MFCs, a much better toxicity tolerance was noticed comparing to DW-MFCs (Fig. 2D). The voltage output was reasonably sustained compared to DW-MFCs, it dropped from 423 mV to 360 mV with ERD concentration increased from 0% to 80%, and the drop was lower than that of DW-MFCs (14.9% vs. 51.1% for AGS-MFCs and DW-MFCs respectively). The variance of the 3 replicates was also less significant than that of the DW-MFCs (2.0% vs. 13.6% for AGS-MFCs and DW-MFCs respectively) at 80% ERD. When 100% of ERD wastewater was injected, the voltage was slightly bounced back with DW-MFCs possibly due to microbial biofilm adapting to the influent chemicals. However, it’s still lower than that of the AGS-MFCs (299 mV for DW-MFCs vs. 341 mV for AGS-MFCs). ERD wastewater, which consists of epichlorohydrin and butylglycidylether, are known to generate biological toxicity towards microorganisms. The compact structure and biologically efficient aerobic granules with diverse microbial species are critical factors for its outstanding toxicity resistance. Aerobic granules covered on top of the biofilm electrode served as an effective shield for the exoelectrogenic biofilm from the attack of the toxic substrate, where the DW-MFC with its anode biofilm exposed unprotectively. This might be the main reason of less voltage degradation and better reproducibility of AGS-MFCs comparing to DW-MFCs. It is possible that if sufficient time was given to DW-MFC to adapt to the target recalcitrant contaminant, the microorganisms would acclimate to build up their toxicity tolerance. However, it usually take a unreasonably long acclimation time. Previous study showed that it took over 100 11
d in forming a microbial system that could degrade sulphamethoxazole to 3-amino-5-methylisoxazole due to the direct inhibition of the antibiotics on microbial activity (Wang et al., 2015), and still the anode biofilm may not be able to compete with the dense and highly functionally packed AGS granules when exposed to shock toxic contaminants. AGS-MFCs also demonstrated superior power generation capabilities when ERD was used as substrate compared to DW-MFCs. No obvious differences on maximum power densities when sodium acetate was used as the substrate, with 408 ± 26 mW/m2 and 404 ± 4 mW/m2 for AGS-MFCs and DW-MFCs respectively (Fig. 3A). Power densities of both types of MFCs were reduced when ERD was used as the substrate, with a 14 % decrease for AGS-MFCs and 29 % decrease for DW-MFCs, due to the toxicity of ERD compounds (Fig. 3A). AGS-MFCs showed max. power densities of 347 ±14 mW/m2, which was 21% higher than that of the DW-MFCs (287 ± 20 mW/m2). Since the electricity generation capabilities of both AGS-MFCs and DW-MFCs using sodium acetate as carbon source demonstrated comparable results, it is assumed that the abundance of exoelectrogens in the anode biofilm would be similar. This result is consistent with previous study using activated sludge and domestic wastewater as inoculums, where no power generation difference was observed when glucose was used as the carbon source (Zhi et al., 2008). Previous study showed that phenol was firstly biodegraded anaerobically to 4-hydroxybenzoic acid and 4-hydroxy-3-methylbenzoic acid, which were then converted to electricity by exoelectrogens accumulated on the 12
anode biofilm (Hedbavna et al., 2016). It indicated the cooperation among different microbial species besides exoelectrogens are necessary for efficient degradation of recalcitrant contaminants. More importantly, the hierarchical structure of AGS is an important factor of the AGS-MFC system being able to withstand the impact of toxins for stable power generation. Polarization curves showed that AGS-MFCs demonstrated slightly higher ohmic resistance than that of the DW-MFC when sodium acetate was used as inoculum (193 Ω with AGS vs. 188 Ω with DW) (Fig. 3B), showing that the AGS granules didn’t affect the internal ohmic resistance significantly. When changing the substrate from sodium acetate to ERD, AGS-MFCs kept the same resistance (193 Ω with sodium acetate vs. 194 Ω with ERD), however, DW-MFCs showed a substantial increase of the internal ohmic resistance (188 Ω with sodium acetate vs. 235 Ω with ERD), which might be another reason leading to the deterioration of the power generation in the DW-MFC systems. Power generation behaviors of both types of MFCs demonstrated that ERD wastewater addition showed obvious impact on bacterial activities, while the AGS-inoculated MFCs were appeared to be affected to a milder extent.
3.3 Wastewater treatment efficiency COD removal was greatly enhanced with AGS inoculated MFCs when ERD was used as the substrate. When sodium acetate was used as the carbon substrate, COD removals of AGS-MFCs were slightly higher than DW-MFCs at the early stage of MFC operation
13
(80.8% and 70.5% at day 15), and they became identical as the operation time proceeded to 25 days and longer (82.1% for AGS-MFCs vs. 80.0% for DW-MFCs, 35 days) (Fig. 4A). As increased concentrations of ERD were used in the influent (from 10% to 80%), the effluent COD of DW-MFCs increased consistently (Fig. 4B), and the effluent COD of DW-MFC reached the maximum of 260.3 mg/L with a lowest COD removal of 51% when ERD was added in the proportion of 80%,. As for AGS-inoculated MFCs, effluent COD increased slightly but recovered in a short time. The highest effluent COD was 230 mg/L with a COD removal of 66%. With 100% ERD addition, the effluent remained to be relatively stable and COD was 118 mg/L with a removal rate of 77%. AGS-MFCs demonstrated better COD removal performance than that of the DW-MFCs. Considering the impact of toxic substance contained in the influent, MFCs inoculated with AGS greatly enhanced their tolerance to the wastewater toxicity. Similar trends were observed for coulombic efficiencies from day 15 to day 25, with AGS-MFCs showing slightly higher CE (26 ± 0.8 % vs. 24 ± 2 %) (Fig. 5). As the operation time proceeded, although identical COD removals were obtained at the same time by both MFC systems, AGS-MFC still showed minor advances over DW-MFCs on CE (28 ± 2 % vs. 26.9 ± 0.3 %).
3.4 Mechanisms of the synergic function of AGS and MFCs A stronger microbial aggregation is a valid defensive strategy against inhibiting effects towards toxic compounds (Jiang et al., 2002). The degradation process of pollutants,
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especially recalcitrant substances usually go through a long chain of decomposition, which is accomplished under the co-functions of multiple types of microbes. For aerobic granular sludge, layers formed at different distances away from granule core are colonized by different microbial species, which was mainly decided by oxygen and nutrients transportation. Anaerobic microorganisms in the central core and facultative microorganisms in the middle layer are not able to live in a substantial number in domestic water and suspended activated sludge. Huge quantity of microbes and rich biodiversity contained in AGS facilitate biofilm formation and enable rapid MFC startup (Fig. S2). Filamentous bacteria interacted with inorganic substance formed the out layer. Bacillius, Grevibacteria formed a microbial net in the middle layer. AGS-attached anode could protect electrogenic microbe from excessive exposures when AGS adsorbed toxins and degraded them into small or less-toxic substrates. In the core of AGS, irregular-shaped microbes could degrade substrate coming from the channel pores in each layer. Bacillus, Brevibacteria, filamentous bacteria were connected in a network by pili, which could serve as nanowire to transfer electrons. Meanwhile, Coccobacteria are tightly attached to the surface of large pylori. The rich and biodiverse microbial community structure is conducive to the operation of MFCs (Du et al., 2007), which is also proved by the minor internal ohmic resistance change by covering the anode electrode. Organic pollutants can be grouped into two categories according to the biotransformation process of microbes: the first category is growth substrate, which 15
provides energy and carbon for cell growth directly when degraded in metabolism. The second category is non-growth substrate, which contributes little or negligible amount of energy or carbon for cell growth. They are generally some pollutants from industrial effluent, hazardous wastes and etc. Non-growth substrate could be degraded in the presence of growth substrate, in which growth substrate serves as both energy provider and electron donor (Bali and Şengül, 2002). Co-substrate is known to affect the simultaneous removal of the target contaminant pentachlorophenol (PCP) in MFCs (Huang et al., 2011). Simultaneous addition of glucose could enhance the system tolerance of PCP shock as well as decrease the acclimation time (Huang et al., 2011). Sodium acetate is one of the most favorable target of microorganisms in wastewater (Kalishwaralal et al., 2009). When the mixture of ERD wastewater and synthesized wastewater (sodium acetate) was used as the influent, sodium acetate facilitated the wastewater treatment by serving as carbon source of microorganisms as well as co-metabolism reagent between electrogenic biofilm and AGS granules. Here in this research, adding sodium acetate as the co-substrate for ERD wastewater treatment didn’t affect the effluent COD in a dramatic way for AGS-MFCs, while a slight overall increase of effluent COD was noticed for DW-MFCs with less sodium acetate (lower effluent COD with 10%-50% ERD vs. higher effluent COD with 50%-100% ERD, Fig. 4B), showing that a higher sodium acetate concentration led to a better final water quality in terms of COD when domestic wastewater was used as the inoculum. Similar power generation, COD removal and coulombic efficiency were obtained for both types 16
of MFCs when only sodium acetate was used as substrate, revealing that biofilm itself was sufficient enough to degrade sodium acetate, a simpler form of carbon source. However, the superior ERD wastewater treatment performance achieved by the AGS-MFCs over DW-MFCs is different from that using sodium acetate as the model substrate (similar results achieved by both systems). This suggests that sodium acetate is not an optimized substrate for MFC performance evaluation to predict the system efficiency when treating more complex wastewaters. Finally, in the AGS-MFC hybrid system, AGS and MFCs could function synergistically to achieve a better performance on both wastewater treatment and bioenergy harvest (Scheme 1). First, AGS with its good adsorption ability could protect anodic biofilm from excessive exposure to toxins in influent (Scheme 1, mechanism ①). AGS is promising for toxic chemical adsorption due to its intrinsic characteristics as high surface area, porosity, hydrophobic property, and rich extracellular polymeric substances. Aerobic granular sludge was prepared as adsorbent to remove heavy metals (Cu 2+) (Adav et al., 2008, Jian et al., 2014). In sewage treatment, the removal of some pharmaceuticals (e.g., diazepam, diclofenac, ibuprofen, naproxen, sulfamethoxazole) was also mainly due to the adsorption of those compounds to sludge. Second, AGS, which serves as a bio-adsorbent in front of biofilm could preliminarily degrade organic compounds into smaller forms of compounds that can be easily decomposed (Scheme 1, mechanism ②). In wastewater treatment, the accessibility of pollutants to microorganisms can be defined in terms of external and internal bioavailability. In this process, physicochemical aspects 17
related to phase distribution and mass transfer, and physiological aspects related to microorganisms such as the presence of active uptake systems, their enzymatic equipment and ability to excrete enzymes, work unitedly on the degradation of pollutants. Microorganisms can transform organic molecules via the succession of oxidation reactions to simpler products (Meer, 2006). Electro-active bacteria attached on the anode then produce electricity through syntrophic interactions through carbon flow (Scheme 1, mechanism③). Further studies of microbial population as well as biomass of the biofilm in a simpler system (e.g. with a single type of carbon source) would be helpful for a better understanding of the synergic effect between AGS and exoelectrogens on recalcitrant organics degradation.
4. Conclusion The addition of AGS covering MFC anode provided for rapid reactor startup as well as efficient treatment of epoxy reactive diluent wastewater. Maximum power density of 491±8 mW/m2 and COD removal of 82±5% were achieved. AGS-MFCs didn't show dramatic advantages on sodium acetate combustion compared to DW-MFCs, however, superior advantages on ERD wastewater treatment were demonstrated due to the combination of AGS’s high toxicity tolerance, its optimized microbial population on biofilm and MFC anode providing readily available electron acceptor. The novel study provided a feasible method for industrial wastewater treatment combining the
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advantages of AGS and MFCs.
Acknowledgements This work was supported by the National Thousand Talents Program, Natural Science Foundation of China (51508213, 51608217 and 21607046), key project of Hubei Provincial Natural Science Foundation (2014CFA109), General program of Natural Science Foundation of Hubei Province (2016CFB539), Innovative and Interdisciplinary Team at HUST (0118261077) and Independent Innovation Foundation of HUST-Exploration Fund (2016YXMS288). The authors would like to thank the Analytical and Testing Center of Huazhong University of Science and Technology for providing the facilities and conduct the characterization work.
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References 1.
Abbasi, U., Jin, W., Pervez, A., Bhatti, Z.A., Tariq, M., Shaheen, S., Iqbal, A., Mahmood, Q., 2016. Anaerobic microbial fuel cell treating combined industrial wastewater: Correlation of electricity generation with pollutants. Bioresour. Technol., 200, 1-7.
2.
Adav, S.S., Lee, D.J., Show, K.Y., Tay, J.H., 2008. Aerobic granular sludge: Recent advances. Biotechnol. Adv., 26, 411-23.
3.
An, B.M., Heo, Y., Maitlo, H.A., Park, J.Y., 2016. Scaled-up dual anode/cathode microbial fuel cell stack for actual ethanolamine wastewater treatment. Bioresour. Technol., 210, 68-73.
4.
Bali, U., Şengül, F., 2002. Performance of a fed-batch reactor treating a wastewater containing 4-chlorophenol. Process Biochem., 37, 1317-1323.
5.
Chaudhuri, S.K., Lovley, D.R., 2003. Electricity generation by direct oxidation of glucose in mediatorless microbial fuel cells. Nature Biotechnol., 21, 1229-32.
6.
Chen T. Aerobic granular sludge stabilization in biocathode chamber of newly constructed continue flow microbial fuel cell system treating synthetic and pharmaceutical wastewater. Desali. Water Treat., 2014:1-10.
7.
Du, Z., Li, H., Gu, T., 2007. A state of the art review on microbial fuel cells: A promising technology for wastewater treatment and bioenergy. Biotechnol. Adv., 25,
20
464-82. 8.
Eaton, A.D., Lenore S.C., Arnold E.G., and Mary Ann H.F., 1998. Standard methods for the examination of water and wastewater. Washington, DC, American Public Health Association.
9.
Feng, Y., Yang, Q., Wang, X., Logan, B.E., 2010. Treatment of carbon fiber brush anodes for improving power generation in air–cathode microbial fuel cells. J Power Sources, 195, 1841-1844.
10. Gao, C.Y., Wang, A.J., Wu, W.M., Yin, Y.L., Zhao, Y.G., 2014. Enrichment of anodic biofilm inoculated with anaerobic or aerobic sludge in single chambered air-cathode microbial fuel cells. Bioresour. Technol., 167C, 124-132. 11. Gao, D., Liu, L., Liang, H., Wu, W., 2011. Aerobic granular sludge: characterization, mechanism of granulation and application to wastewater treatment. Crit. Rev. Biotechnol., 31, 137-152. 12. Habibul, N., Hu, Y., Wang, Y., Chen, W., Yu, H., Sheng, G., 2016. Bioelectrochemical Chromium (VI) Removal in Plant-Microbial Fuel Cells. Environ. Sci. Technol., 50, 3882-3889. 13. Hou, H., Li, L., Cho, Y., De Figueiredo, P., Han, A., 2009. Microfabricated microbial fuel cell arrays reveal electrochemically active microbes. PLOS one, 4, e6570. 14. Huang, L., Gan, L., Zhao, Q., Logan, B.E., Lu, H., Chen, G., 2011. Degradation of pentachlorophenol with the presence of fermentable and non-fermentable 21
co-substrates in a microbial fuel cell. Bioresour. Technol., 102, 8762-8. 15. Jian M, Tang C, Liu M. Adsorptive removal of Cu2+ from aqueous solution using aerobic granular sludge. Desali. Water Treat., 2014, 54(7):1-10. 16. Jiang, H.L., Tay, J.H., Tay, T.L., 2002. Aggregation of immobilized activated sludge cells into aerobically grown microbial granules for the aerobic biodegradation of phenol. Lett. Appl. Microb. 35, 439-45. 17. Kalishwaralal, K., Deepak, V., Pandian, S.R.K., Gurunathan, S., 2009. Biological synthesis of gold nanocubes from Bacillus licheniformis. Bioresour. Technol., 100, 5356-5358. 18. Kiely, P.D., Cusick, R., Call, D.F., Selembo, P.A., Regan, J.M., Logan, B.E., 2011. Anode microbial communities produced by changing from microbial fuel cell to microbial electrolysis cell operation using two different wastewaters. Bioresour. Technol., 102, 388-394. 19. Kim, J.R., Premier, G.C., Hawkes, F.R., Rodríguez, J., Dinsdale, R.M., Guwy, A.J., 2010. Modular tubular microbial fuel cells for energy recovery during sucrose wastewater treatment at low organic loading rate. Bioresour. Technol., 101, 1190-1198. 20. Kim, K.Y., Yang, W., Logan, B.E., 2015. Impact of electrode configurations on retention time and domestic wastewater treatment efficiency using microbial fuel cells. Water Res., 80, 41-46. 22
21. Liu, B., Li, B., 2014. Single chamber microbial fuel cells (SCMFCs) treating wastewater containing methanol. Int. J Hydrogen Energ., 39, 2340-2344. 22. Liu, G., Yates, M.D., Cheng, S., Call, D.F., Sun, D., Logan, B.E., 2011. Examination of microbial fuel cell start-up times with domestic wastewater and additional amendments. Bioresour. Technol., 102, 7301-6. 23. Liu, H., Logan, B.E., 2004. Electricity generation using an air-cathode single chamber microbial fuel cell in the presence and absence of a proton exchange membrane. Environ. Sci. Technol., 38, 4040-6. 24. Liu, H., Ramnarayanan, R., Logan, B.E., 2004. Production of electricity during wastewater treatment using a single chamber microbial fuel cell. Environ. Sci. Technol., 38, 2281-2285. 25. Logan, B.E., Hamelers, B., Rozendal, R., SchrÖder, U., Keller, J., Freguia, S, Aelterman, P., Verstraete, W., Rabaey, K., 2006. Microbial fuel cells: Methodology and Technology.
Environ. Sci. Technol., 40, 5181-5192
26. Long, B., Yang, C., Pu, W., Yang, J., Jiang, G., Dan, J., Li, C., Liu, F., 2014. Rapid cultivation of aerobic granular sludge in a pilot scale sequencing batch reactor. Bioresour. Technol., 166, 57-63. 27. Lovley, D.R., Phillips, E.J., 1988. Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese. Appl. Environ. Microb., 54, 1472-1480. 23
28. Luo, H., Liu, G., Zhang, R., Jin, S., 2009. Phenol degradation in microbial fuel cells. Chem. Eng. J., 147, 259-264. 29. Meer, J.R.V.D., 2006. Environmental pollution promotes selection of microbial degradation pathways. Front. Ecol. Environ., 4, 35-42. 30. Miran, W., Nawaz, M., Kadam, A., Shin, S., Heo, J., Jang, J., Lee, D.S., 2015. Microbial community structure in a dual chamber microbial fuel cell fed with brewery waste for azo dye degradation and electricity generation. Environ. Sci. Pollut. R., 22, 13477-13485. 31. Ogugbue, C., Ebode, E., Leera, S., 2015. Electricity generation from swine wastewater using microbial fuel cell., Water Res., 16, 26-33. 32. Pant, D., Bogaert, G.V., Smet, M.D., Diels, L., Vanbroekhoven, K., 2010. Use of novel permeable membrane and air cathodes in acetate microbial fuel cells. Electrochim. Acta, 55, 7710-7716. 33. Pant, D., Van Bogaert, G., Diels, L., Vanbroekhoven, K., 2010. A review of the substrates used in microbial fuel cells (MFCs) for sustainable energy production. Bioresour. Technol., 101, 1533-1543. 34. Rinaldi, A., Mecheri, B., Garavaglia, V., Licoccia, S., Di Nardo, P., Traversa, E., 2008. Engineering materials and biology to boost performance of microbial fuel cells: a critical review. Energ. Environ. Sci., 1, 417.
24
35. Sahinkaya, E., Dilek, F.B., 2005. Biodegradation of 4-chlorophenol by acclimated and unacclimated activated sludge—Evaluation of biokinetic coefficients. Environ. Res., 99, 243-252. 36. Sonawane, J.M., Marsili, E., Ghosh, P.C., 2014. Treatment of domestic and distillery wastewater in high surface microbial fuel cells. Int. J. Hydrogen Energ., 39, 21819-21827. 37. Wang, C., Deng, H., Zhao, F., 2016. The remediation of chromium (VI)-contaminated soils using microbial fuel cells. Soil Sediment Contaminat., 25, 1-12. 38. Wen, Q., Ying, W., Cao, D., Zhao, L., Qian, S., 2009. Electricity generation and modeling of microbial fuel cell from continuous beer brewery wastewater. Bioresour. Technol., 100, 4171-4175. 39. Yang, W., He, W., Zhang, F., Hickner, M.A., Logan, B.E., 2014. Single-Step Fabrication Using a Phase Inversion Method of Poly (vinylidene fluoride) (PVDF) Activated Carbon Air Cathodes for Microbial Fuel Cells. Environ. Sci. Technol. Lett., 1, 416-420. 40. Yanuka-Golub, K., Reshef, L., Rishpon, J., Gophna, U., 2016. Community structure dynamics during startup in microbial fuel cells – the effect of phosphate concentrations. Bioresour. Technol., 212, 151-159. 41. Zhang, F., Xia, X., Luo, Y., Sun, D., Call, D.F., Logan, B.E., 2013. Improving startup 25
performance with carbon mesh anodes in separator electrode assembly microbial fuel cells. Bioresour. Technol., 133, 74-81. 42. Zhi, W., Ge, Z., He, Z., Zhang, H., 2014. Methods for understanding microbial community structures and functions in microbial fuel cells: a review. Bioresour. Technol., 171, 461-468. 43. Zhi, Y., Liu, H., Yao, L., 2008. The Effect of Suspended Sludge on Electricity Generation in Microbial Fuel Cells. Int. Conf. Bioinform. Biomed. Eng., pp. 2923-2927
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Figure Captions Fig. 1 Setup of the AGS-MFC system: a dense layer of AGS covering the MFC anode electrode. The AGS layer is held in place with a stainless steel mesh;
Fig. 2 (A) Startup of both DW-MFCs and AGS-MFCs; (B) Stable power generation of AGS-MFCs using 1 g/L sodium acetate as carbon source; (C) Voltage generation of DW-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate; (D) Voltage generation of AGS-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate.
Fig. 3 (A) Power curves and (B) polarization curves of DW/AGS-MFCs treating different types of wastewater (Sodium acetate-NaAc and ERD wastewater-ERD).
Fig. 4 COD concentrations in the influent and the effluent and COD removal of DW-MFCs and AGS-MFCs treating different types of wastewater: (A) synthesized waster (1 g/L sodium acetate) and (B) ERD wastewater with an increasing proportion.
Fig. 5 Coulombic efficiencies of DW-MFCs and AGS-MFCs treating different types of wastewater. (A) ERD wastewater and (B) synthesized waster (1 g/L sodium acetate).
Scheme 1 Proposed mechanisms of rapid AGS-MFC startup by facilitating biofilm formation as well as synergetic biodegradation of toxins by MFC biofilm anode with AGS granules during ERD wastewater treatment (①: AGS adsorbs a certain amount of
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toxins in the solution, forming a protective layer to prohibit the direct attack of toxins towards the biofilm anode; ②: AGS degrades toxins into more biodegradable components that could penetrate the AGS layer to the biofilm surface; ③: Bioelectricity generation by electrogenic biofilm anode converting carbon source to electrons).
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Figures
Fig. 1 Setup of the AGS-MFC system: a dense layer of AGS covering the MFC anode electrode. The AGS layer is held in place with a stainless steel mesh;
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Fig. 2 (A) Startup of both DW-MFCs and AGS-MFCs; (B) Stable power generation of AGS-MFCs using 1 g/L sodium acetate as carbon source; (C) Voltage generation of DW-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate; (D) Voltage generation of AGS-MFCs with increased concentrations from 10% to 100% of ERD wastewater mixed with 1 g/L sodium acetate.
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B 600
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Fig. 3 (A) Power curves and (B) polarization curves of DW/AGS-MFCs treating different types of wastewater (Sodium acetate-NaAc and ERD wastewater-ERD).
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Fig. 4 COD concentrations in the influent and the effluent and COD removal of DW-MFCs and AGS-MFCs treating different types of wastewater: (A) synthesized waster (1 g/L sodium acetate) and (B) ERD wastewater with an increasing proportion.
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Fig. 5 Coulombic efficiencies of DW-MFCs and AGS-MFCs treating different types of wastewater. (A) ERD wastewater and (B) synthesized waster (1 g/L sodium acetate).
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Scheme 1 Proposed mechanisms of rapid AGS-MFC startup by facilitating biofilm formation as well as synergetic biodegradation of toxins by MFC biofilm anode with AGS granules during ERD wastewater treatment (①: AGS adsorbs a certain amount of toxins in the solution, forming a protective layer to prohibit the direct attack of toxins towards the biofilm anode; ②: AGS degrades toxins into more biodegradable components that could penetrate the AGS layer to the biofilm surface; ③: Bioelectricity generation by electrogenic biofilm anode converting carbon source to electrons).
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Graphical Abstract
35
Highlights •
Aerobic granular sludge was incorporated into MFCs to form a hybrid system.
•
AGS-MFCs shortened the startup time from 13 d to 7 d compared to MFCs without
AGS. •
Superior performance for epoxy reactive diluent wastewater treatment was achieved.
•
The high efficiency is possibly due to the synergic effect of both MFC and AGS.
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