Chemosphere 117 (2014) 151–157
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Integration of microbial fuel cell techniques into activated sludge wastewater treatment processes to improve nitrogen removal and reduce sludge production Shashikanth Gajaraj, Zhiqiang Hu ⇑ Department of Civil and Environmental Engineering, University of Missouri, Columbia, MO 65211, United States
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
Microbial fuel cell technique was
Air
incorporated into the MLE and MBR process. No changes in sludge activity, settling property, and nitrifying community structure. Effluent COD concentration of the MLE–MFC was lower than that of MLE. MLE–MFC and MBR–MFC increased NO3–N removal by 31% and 20% respectively. The MFC integrated systems reduced sludge production by 6–11%.
Influent
a r t i c l e
a b s t r a c t
i n f o
Article history: Received 23 October 2013 Received in revised form 23 May 2014 Accepted 5 June 2014
Handling Editor: Chang-Ping Yu Keywords: Microbial fuel cell Modified Ludzack–Ettinger process Membrane bioreactor Bioelectrochemical systems Nitrogen removal Sludge reduction
Effluent
1. Graphite cloth 2. Ti mesh 3. Conductive foam Anodic Chamber
Bioelectrochemical systems are emerging for wastewater treatment, yet little is known about how well they can be integrated with current wastewater treatment processes. In this bench-scale study, the microbial fuel cell (MFC) technique was incorporated into the Modified Ludzack–Ettinger (MLE) process (phase I) and later with the membrane bioreactor (MBR) process (phase II) to evaluate the performance of MFC assisted wastewater treatment systems (i.e., MLE–MFC and MBR–MFC). There was no significant difference in the effluent NH+4–N concentration between the systems integrating MFC and the open circuit controls. The average effluent COD concentration was significantly lower in the MLE–MFC, but it did not change much in the MBR–MFC because of the already low COD concentrations in MBR operation. The MLE–MFC and MBR–MFC systems increased the NO 3 –N removal efficiencies by 31% (±12%) and 20% (±12%), respectively, and reduced sludge production by 11% and 6%, respectively, while generating an average voltage of 0.13 (±0.03) V in both systems. Analysis of the bacterial specific oxygen uptake rate, the sludge volume index, and ammonia-oxidizing bacterial population (dominated by Nitrosomonas through terminal restriction fragment length polymorphism analysis) indicated that there was no significant difference in sludge activity, settling property, and nitrifying community structure between the MFC assisted systems and the open circuit controls. The results suggest that the wastewater treatment systems could achieve higher effluent water quality and lower sludge production if it is integrated well with MFC techniques. Ó 2014 Elsevier Ltd. All rights reserved.
⇑ Corresponding author. Tel.: +1 573 884 0497; fax: +1 573 882 4784. E-mail address: [email protected] (Z. Hu). http://dx.doi.org/10.1016/j.chemosphere.2014.06.013 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.
Cathodic Chamber
152
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
1. Introduction Bioelectrochemical systems (BES) are emerging for wastewater treatment and energy production (Rozendal et al., 2008). Microbial fuel cells (MFC) are the archetype of BES and have been demonstrated successful in wastewater treatment (Ahn and Logan, 2013; Kim et al., 2013) although the power output is not high enough for the electricity production process (Logan and Rabaey, 2012). Research on MFC has mainly been focused on increasing power densities (Logan, 2009) although the ability to scale-up the process remains to be determined (Dewan et al., 2008). Numerous designs have been proposed to optimize MFC operations. There are two chamber systems, single chamber systems, up-flow mode systems, and stacked MFC systems (Du et al., 2007). The similarity in all these designs points to the presence of an anaerobic zone and an aerobic zone. Indeed, enhanced biological nutrient removal processes such as the Modified Ludzack–Ettinger (MLE) process, rely on the use of alternating aerobic and anoxic conditions (Yu et al., 2011). Furthermore, membrane bioreactor (MBR) technology can be used in these nutrient removal wastewater treatment plants (WWTPs) to produce higher quality effluent (Ge et al., 2013). By integrating MFC technology with these common wastewater treatment processes, many benefits could be gained including electricity production, lower sludge production (Logan et al., 2006) and therefore reduced sludge handling costs. Taking MBR activated sludge process as an example, it was demonstrated that a small electric field (0.036–0.073 V cm1) mitigated fouling of conductive membranes in the MBR (Liu et al., 2012b) but required novel materials for the membrane construction. Wang et al. (2011) and Ge et al. (2013) used a stainless steel mesh of 40 lm pore size and hollow fiber membrane of 0.02 lm pore size membrane respectively as the cathode for sludge filtration and higher effluent water quality. However these designs had to bear the cost associated with the use of cation exchange membrane (CEM), had limited nutrient removal due to the reactor design, and furthermore, the effect of bio-generated electricity in mitigating membrane fouling was not studied. A slightly modified design (Kim et al., 2013) used an ultrafiltration membrane with a molecular weight cutoff of 1 kDa instead of a CEM. Although the cost associated with the use of CEM was avoided, the design called for a positive pressure system which posed scale-up issues and no nutrients were removed. While all the aforementioned demonstrated integrating MBR with MFC processes at lab scale they fall short of immediate potential for successful scale-up or integration with existing WWTPs. One more important benefit in the MFC systems to be explored is the enhanced nitrogen removal from wastewater (Virdis et al., 2010; Yu et al., 2011; Zhang and Angelidaki, 2012). In one study, MFC with a reactor volume of 0.336 L was coupled with an external aerobic nitrification reactor to convert ammonia in the feed solution to nitrate before it was circulated through the MFC cathodic chamber where bacteria on biocathode play a role for nitrogen removal (Virdis et al., 2008). However, the set-up of an additional external nitrifying bioreactor makes it difficult to integrate with existing wastewater treatment systems. Furthermore, the cost associated with proton exchange membrane (PEM) and high recirculation flow to support in-situ nitrification (Clauwaert et al., 2007; Virdis et al., 2010) can often be prohibitive for wastewater treatment. Similarly, the complexity of the dual-cathode MFCs would not allow to easily integrate MFC techniques into traditional wastewater treatment operations, although such a system was capable of generating electricity and removing nitrogen more efficiently (Zhang and He, 2012). The objective of this study was to evaluate the performance of MFC assisted wastewater treatment systems, to improve
wastewater treatment and sludge management practices. Unlike the previous systems where small standalone MFC modules were used in sequence (Kim et al., 2013) or immersed in bioreactors (Zhang and Angelidaki, 2012; Ahn and Logan, 2013), we operated the MFC assisted activated sludge wastewater treatment systems (7.2 L each) in consistence with the use of alternating aerobic and anoxic conditions in the MLE or MBR process. These processes were selected because MLE is one of the most commonly used processes for enhanced nitrogen removal (USEPA, 2000) and MBR becomes increasingly used for higher effluent water quality and water reuse applications.
2. Materials and methods 2.1. Bioreactor design and operation Two identical bioreactors were constructed with glass, each having a working volume of 7.2 L. A schematic of the bioreactors and the flow pattern is shown in Fig. 1. The bioreactors were divided into three zones: anaerobic/anoxic chamber (far left), aerobic chamber (middle) and an internal settling chamber (far right) separated by plastic baffles. The effective volumes of the resulting chambers were 2.1, 3.3, and 1.8 L, respectively. An array of three horizontal openings, each 0.5 cm in diameter was made in the baffle wall separating the anaerobic and aerobic chamber. An MFC module was integrated into one of the bioreactors by placing the anode in the anaerobic/anoxic chamber and the cathode in the aerobic chamber (Fig. 1). The other bioreactor served as the control, which included a similar electrode module, but operated on open circuit. The anode was made of 120 cm2 (8 cm 15 cm) of commercially available graphite cloth (Plantraco, Saskatoon, SK, Canada). The cloth was in plain weave (checkerboard pattern) and had a specific weight of 0.08 kg m3 with a resistivity of 1 105 O m. The cathode was designed to be a hollow cylinder so that it could accommodate the membrane module during the MBR study (Fig. 1). The hollow cylindrical core was made out of a rigid polypropylene mesh tube (catalog number RN1900), 15 cm in length, 6 cm in diameter and an open area of 35% (Industrial netting, Minneapolis, MN, USA). The rigid hollow tube was covered with a single layer of the following materials in the following order: (1) Carbon fiber/graphite cloth, which served as a primary electrode material. (2) Titanium mesh of 0.28 mm diameter wire in a 12 12 wires per 2.54 cm plain weave (Unique wire weaving, Hillside, NJ, USA) for electrical connection, and (3) Conductive low-density black polyurethane foam (All-spec industries, Wilmington, NC, USA), which served as a primary filter to support biofilm growth on its large surface area and control membrane fouling in the MBR study. To construct the MFC module, the carbon fiber/graphite cloth and the titanium mesh were soaked in 50% ethanol for about 1 h and rinsed with tap water to wash away any impurities before use. The anode and cathode were connected together with a 1000 X resistor and spaced approximately 5 cm apart from each side of the baffle plate. PEM was not used due to cost constraints. The bioreactors were inoculated with return activated sludge from the Columbia Wastewater Treatment Plant (Columbia, MO, USA). The reactors were fed with synthetic wastewater that was mainly composed of nonfat milk powder with a COD of approximately 500 mg L1 (Liu et al., 2012b). Other components of the synthetic wastewater included 51.7 mg L1 total nitrogen (TN),
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
153
Fig. 1. Schematic of conventional (open circuit control) and MFC-integrated wastewater treatment systems (MLE–MFC in phase I and MBR–MFC in phase II). Black arrows denote direction of liquid flow.
30 mg L1 ammonia–nitrogen (NH+4–N), and 6 mg L1 total phosphorus (TP). The synthetic wastewater also contained trace metals in the following concentrations (in mg L1): 44 MgSO4, 15 CaCl22H2O, 2 FeCl24H2O, 3.4 MnSO4H2O, 1.2 (NH4)6Mo7O244H2O, 0.8 CuSO4, and 1.8 Zn(NO3)2H2O. The wastewater treatment systems were operated at hydraulic retention time of 1 d at a flow rate of 7.2 L d1 and a target solids retention time (SRT) of 20 d. A mixed liquor recirculation (sludge recycling) ratio of 1:1 (mixed liquor to influent flow) was maintained in all the systems. The mixed liquor was re-circulated from the settling zone back into the anaerobic zone (Fig. 1). The study was divided into two phases. During phase I (from day 30 following acclimatization to day 100) the bioreactor was run in the MLE mode. The influent was fed into the anoxic zone and the mixed liquor was returned from the settling zone to the anoxic zone (Fig. 1). A fine bubble stone diffuser adjacent to the cathode, 15 cm away from the anode provided aeration in the aerobic chamber. At the end of phase I, sludge wasted was reduced by half the normal rate for 14 d for biomass recovery, following which the sludge from both MLE systems were mixed completely to ascertain the same initial biomass concentration at the beginning of the MBR study. Once the MBR was fully operational, sludge wastage was again increased to maintain the target SRT of 20 days. During phase II (from day 100 to day 210) a membrane module was added to the aerobic zone so that the bioreactor was run as a MBR. The membrane was inserted inside the hollow cathode in the MBR–MFC while it was placed besides the open circuit cathode in the control run to demonstrate the unique filtering feature of the cathode (Fig. 1). The submerged membrane module (ZeeWeed hollow fiber membrane module, GE Water & Process Technologies, Trevose, PA, USA) was made of polyvinylidene fluoride with a nominal pore size of 0.1 lm, a total effective filtration area of 0.047 m2, and a designed membrane flux of 15.0 L m2 h1. The permeate
pumps operating on variable drives were adjusted regularly to maintain a constant flux of 4.8 L m2 h1. An air pump supplied compressed air to the built-in orifices at the bottom of each membrane module at a constant air flow rate of 6 L min1 to support aeration and control membrane fouling. Throughout the study period, magnetic stirrers were used to provide additional mixing of the mixed liquor in the anoxic and oxic chambers. 2.2. Sludge activity measurements Activated sludge activities were assessed with the specific oxygen uptake rate (SOUR) measurements in duplicate samples. Briefly, aliquots of sludge collected from the aerobic chamber in each bioreactor were aerated with pure oxygen and transferred to the respirometric bottles, which were tightly capped after insertion of a DO probe in each bottle with no headspace (air). At a predetermined time between 500 and 600 s, an aliquot of substrate (10 mg L1 NH+4–N or 20 mg COD L1 of acetate) was injected with a 10 lL glass syringe. The decrease in DO concentration in the respirometric bottle due to substrate oxidation was measured by a DO monitor (YSI model 5300A, Yellow Springs, OH, USA) and continuously recorded at 4 Hz by LabView software interfaced with a computer. The SOUR was calculated by linear regression analysis. 2.3. Membrane flux monitoring and fouling control The trans-membrane pressure (TMP) and permeate flux of the membrane module in each MBR (i.e., MBR–MFC and the open circuit control) were closed monitored using an inline battery-powered digital pressure gauge (model EW-68935-18, Cole-Parmer, Vernon Hills, IL, USA) during the phase II study. The membrane flux was maintained relatively constant at 4.8 L m2 h1 by adjusting the speed of the permeate pump. When the TMP increased and
154
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
exceeded 45 kPa (maximum allowable value of 50 kPa as suggested by the manufacturer), the membrane module was taken out of the MBR for physical cleaning. The membrane module was rinsed with approximately 30 L of distilled water under pressure to remove the attached biofilm. The membrane module was then soaked in distilled water for 0.5 h before it was submerged back to the bioreactor. No chemical cleaning was performed. 2.4. Analytical methods and statistical analysis Influent and effluent water quality was monitored twice a week during the study period. The wastewater effluent (phase I) and the permeate water (phase II) parameters such as NH+4–N, NO 3 –N, NO 2 –N, and COD were analyzed according to the standard methods (APHA, 1998). The biomass concentration, sludge volume index (SVI) in the MLE systems, and MBRs were measured in duplicate (APHA, 1998). Voltage generated by the MFC assisted system was monitored in real-time at 1 Hz using LabView data acquisition system (National Instruments, TX, USA) and the voltage was averaged every 24 h. Coulombic efficiency (gc) defined as the ratio of electrons used as current to the maximum electron production (Logan et al., 2006) was calculated as follows:
gc ð%Þ ¼
MI 8I ¼ FbqDCOD FqDCOD
where M = 32 (mol. wt. of oxygen), I is the current (A), F is Faradays constant, b = 4 (the number of electrons exchanged per mol of oxygen), q is the flow rate (L s1), and DCOD is the difference between the influent and effluent COD. The analysis of variance (ANOVA) was applied to compare the means of different groups of the study where p values less than 0.05 were considered statistically significant. 2.5. Nitrifying bacterial community analysis Terminal restriction fragment length polymorphism (T-RFLP) was used to analyze and compare the effect of MFC assisted wastewater treatment systems on the nitrifying community structure (nitrification potential) between the MLE–MFC in phase I and MBR–MFC in phase II and the open circuit control, based on the known 16S rRNA genes of ammonia-oxidizing bacteria (AOB) as described previously (Siripong and Rittmann, 2007). Activated sludge samples were collected from the aerobic chamber in the MLE and MBR systems on day 80 and day 160 respectively. Detailed information about DNA extraction and T-RFLP analysis is described in Supporting Information.
Fig. 2. Voltage output from the MLE–MFC (phase I) and MBR–MFC (phase II) over the study period.
NH+4–N. Furthermore, no nitrites were detected in the effluent. The effluent NO 3 –N concentrations in the MLE–MFC (phase I) and MBR–MFC system (phase II) were 4.7 ± 1.3 and 6.2 ± 1.5 mg L1, respectively. By contrast, the NO 3 –N concentrations in the open circuit controls were 6.9 ± 1.8 and 7.8 ± 1.5 mg L1, respectively (Fig. 3). There was no significant difference in effluent nitrate concentration between the regular MLE and MBR processes (p = 0.22). However, regardless of the type of process used in wastewater treatment, the effluent NO 3 –N concentrations of the MLE–MFC and MBR–MFC were significantly lower than those of the controls (p values
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Integration of microbial fuel cell techniques into activated sludge wastewater treatment processes to improve nitrogen removal and reduce sludge production Shashikanth Gajaraj, Zhiqiang Hu ⇑ Department of Civil and Environmental Engineering, University of Missouri, Columbia, MO 65211, United States
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
Microbial fuel cell technique was
Air
incorporated into the MLE and MBR process. No changes in sludge activity, settling property, and nitrifying community structure. Effluent COD concentration of the MLE–MFC was lower than that of MLE. MLE–MFC and MBR–MFC increased NO3–N removal by 31% and 20% respectively. The MFC integrated systems reduced sludge production by 6–11%.
Influent
a r t i c l e
a b s t r a c t
i n f o
Article history: Received 23 October 2013 Received in revised form 23 May 2014 Accepted 5 June 2014
Handling Editor: Chang-Ping Yu Keywords: Microbial fuel cell Modified Ludzack–Ettinger process Membrane bioreactor Bioelectrochemical systems Nitrogen removal Sludge reduction
Effluent
1. Graphite cloth 2. Ti mesh 3. Conductive foam Anodic Chamber
Bioelectrochemical systems are emerging for wastewater treatment, yet little is known about how well they can be integrated with current wastewater treatment processes. In this bench-scale study, the microbial fuel cell (MFC) technique was incorporated into the Modified Ludzack–Ettinger (MLE) process (phase I) and later with the membrane bioreactor (MBR) process (phase II) to evaluate the performance of MFC assisted wastewater treatment systems (i.e., MLE–MFC and MBR–MFC). There was no significant difference in the effluent NH+4–N concentration between the systems integrating MFC and the open circuit controls. The average effluent COD concentration was significantly lower in the MLE–MFC, but it did not change much in the MBR–MFC because of the already low COD concentrations in MBR operation. The MLE–MFC and MBR–MFC systems increased the NO 3 –N removal efficiencies by 31% (±12%) and 20% (±12%), respectively, and reduced sludge production by 11% and 6%, respectively, while generating an average voltage of 0.13 (±0.03) V in both systems. Analysis of the bacterial specific oxygen uptake rate, the sludge volume index, and ammonia-oxidizing bacterial population (dominated by Nitrosomonas through terminal restriction fragment length polymorphism analysis) indicated that there was no significant difference in sludge activity, settling property, and nitrifying community structure between the MFC assisted systems and the open circuit controls. The results suggest that the wastewater treatment systems could achieve higher effluent water quality and lower sludge production if it is integrated well with MFC techniques. Ó 2014 Elsevier Ltd. All rights reserved.
⇑ Corresponding author. Tel.: +1 573 884 0497; fax: +1 573 882 4784. E-mail address: [email protected] (Z. Hu). http://dx.doi.org/10.1016/j.chemosphere.2014.06.013 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.
Cathodic Chamber
152
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
1. Introduction Bioelectrochemical systems (BES) are emerging for wastewater treatment and energy production (Rozendal et al., 2008). Microbial fuel cells (MFC) are the archetype of BES and have been demonstrated successful in wastewater treatment (Ahn and Logan, 2013; Kim et al., 2013) although the power output is not high enough for the electricity production process (Logan and Rabaey, 2012). Research on MFC has mainly been focused on increasing power densities (Logan, 2009) although the ability to scale-up the process remains to be determined (Dewan et al., 2008). Numerous designs have been proposed to optimize MFC operations. There are two chamber systems, single chamber systems, up-flow mode systems, and stacked MFC systems (Du et al., 2007). The similarity in all these designs points to the presence of an anaerobic zone and an aerobic zone. Indeed, enhanced biological nutrient removal processes such as the Modified Ludzack–Ettinger (MLE) process, rely on the use of alternating aerobic and anoxic conditions (Yu et al., 2011). Furthermore, membrane bioreactor (MBR) technology can be used in these nutrient removal wastewater treatment plants (WWTPs) to produce higher quality effluent (Ge et al., 2013). By integrating MFC technology with these common wastewater treatment processes, many benefits could be gained including electricity production, lower sludge production (Logan et al., 2006) and therefore reduced sludge handling costs. Taking MBR activated sludge process as an example, it was demonstrated that a small electric field (0.036–0.073 V cm1) mitigated fouling of conductive membranes in the MBR (Liu et al., 2012b) but required novel materials for the membrane construction. Wang et al. (2011) and Ge et al. (2013) used a stainless steel mesh of 40 lm pore size and hollow fiber membrane of 0.02 lm pore size membrane respectively as the cathode for sludge filtration and higher effluent water quality. However these designs had to bear the cost associated with the use of cation exchange membrane (CEM), had limited nutrient removal due to the reactor design, and furthermore, the effect of bio-generated electricity in mitigating membrane fouling was not studied. A slightly modified design (Kim et al., 2013) used an ultrafiltration membrane with a molecular weight cutoff of 1 kDa instead of a CEM. Although the cost associated with the use of CEM was avoided, the design called for a positive pressure system which posed scale-up issues and no nutrients were removed. While all the aforementioned demonstrated integrating MBR with MFC processes at lab scale they fall short of immediate potential for successful scale-up or integration with existing WWTPs. One more important benefit in the MFC systems to be explored is the enhanced nitrogen removal from wastewater (Virdis et al., 2010; Yu et al., 2011; Zhang and Angelidaki, 2012). In one study, MFC with a reactor volume of 0.336 L was coupled with an external aerobic nitrification reactor to convert ammonia in the feed solution to nitrate before it was circulated through the MFC cathodic chamber where bacteria on biocathode play a role for nitrogen removal (Virdis et al., 2008). However, the set-up of an additional external nitrifying bioreactor makes it difficult to integrate with existing wastewater treatment systems. Furthermore, the cost associated with proton exchange membrane (PEM) and high recirculation flow to support in-situ nitrification (Clauwaert et al., 2007; Virdis et al., 2010) can often be prohibitive for wastewater treatment. Similarly, the complexity of the dual-cathode MFCs would not allow to easily integrate MFC techniques into traditional wastewater treatment operations, although such a system was capable of generating electricity and removing nitrogen more efficiently (Zhang and He, 2012). The objective of this study was to evaluate the performance of MFC assisted wastewater treatment systems, to improve
wastewater treatment and sludge management practices. Unlike the previous systems where small standalone MFC modules were used in sequence (Kim et al., 2013) or immersed in bioreactors (Zhang and Angelidaki, 2012; Ahn and Logan, 2013), we operated the MFC assisted activated sludge wastewater treatment systems (7.2 L each) in consistence with the use of alternating aerobic and anoxic conditions in the MLE or MBR process. These processes were selected because MLE is one of the most commonly used processes for enhanced nitrogen removal (USEPA, 2000) and MBR becomes increasingly used for higher effluent water quality and water reuse applications.
2. Materials and methods 2.1. Bioreactor design and operation Two identical bioreactors were constructed with glass, each having a working volume of 7.2 L. A schematic of the bioreactors and the flow pattern is shown in Fig. 1. The bioreactors were divided into three zones: anaerobic/anoxic chamber (far left), aerobic chamber (middle) and an internal settling chamber (far right) separated by plastic baffles. The effective volumes of the resulting chambers were 2.1, 3.3, and 1.8 L, respectively. An array of three horizontal openings, each 0.5 cm in diameter was made in the baffle wall separating the anaerobic and aerobic chamber. An MFC module was integrated into one of the bioreactors by placing the anode in the anaerobic/anoxic chamber and the cathode in the aerobic chamber (Fig. 1). The other bioreactor served as the control, which included a similar electrode module, but operated on open circuit. The anode was made of 120 cm2 (8 cm 15 cm) of commercially available graphite cloth (Plantraco, Saskatoon, SK, Canada). The cloth was in plain weave (checkerboard pattern) and had a specific weight of 0.08 kg m3 with a resistivity of 1 105 O m. The cathode was designed to be a hollow cylinder so that it could accommodate the membrane module during the MBR study (Fig. 1). The hollow cylindrical core was made out of a rigid polypropylene mesh tube (catalog number RN1900), 15 cm in length, 6 cm in diameter and an open area of 35% (Industrial netting, Minneapolis, MN, USA). The rigid hollow tube was covered with a single layer of the following materials in the following order: (1) Carbon fiber/graphite cloth, which served as a primary electrode material. (2) Titanium mesh of 0.28 mm diameter wire in a 12 12 wires per 2.54 cm plain weave (Unique wire weaving, Hillside, NJ, USA) for electrical connection, and (3) Conductive low-density black polyurethane foam (All-spec industries, Wilmington, NC, USA), which served as a primary filter to support biofilm growth on its large surface area and control membrane fouling in the MBR study. To construct the MFC module, the carbon fiber/graphite cloth and the titanium mesh were soaked in 50% ethanol for about 1 h and rinsed with tap water to wash away any impurities before use. The anode and cathode were connected together with a 1000 X resistor and spaced approximately 5 cm apart from each side of the baffle plate. PEM was not used due to cost constraints. The bioreactors were inoculated with return activated sludge from the Columbia Wastewater Treatment Plant (Columbia, MO, USA). The reactors were fed with synthetic wastewater that was mainly composed of nonfat milk powder with a COD of approximately 500 mg L1 (Liu et al., 2012b). Other components of the synthetic wastewater included 51.7 mg L1 total nitrogen (TN),
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
153
Fig. 1. Schematic of conventional (open circuit control) and MFC-integrated wastewater treatment systems (MLE–MFC in phase I and MBR–MFC in phase II). Black arrows denote direction of liquid flow.
30 mg L1 ammonia–nitrogen (NH+4–N), and 6 mg L1 total phosphorus (TP). The synthetic wastewater also contained trace metals in the following concentrations (in mg L1): 44 MgSO4, 15 CaCl22H2O, 2 FeCl24H2O, 3.4 MnSO4H2O, 1.2 (NH4)6Mo7O244H2O, 0.8 CuSO4, and 1.8 Zn(NO3)2H2O. The wastewater treatment systems were operated at hydraulic retention time of 1 d at a flow rate of 7.2 L d1 and a target solids retention time (SRT) of 20 d. A mixed liquor recirculation (sludge recycling) ratio of 1:1 (mixed liquor to influent flow) was maintained in all the systems. The mixed liquor was re-circulated from the settling zone back into the anaerobic zone (Fig. 1). The study was divided into two phases. During phase I (from day 30 following acclimatization to day 100) the bioreactor was run in the MLE mode. The influent was fed into the anoxic zone and the mixed liquor was returned from the settling zone to the anoxic zone (Fig. 1). A fine bubble stone diffuser adjacent to the cathode, 15 cm away from the anode provided aeration in the aerobic chamber. At the end of phase I, sludge wasted was reduced by half the normal rate for 14 d for biomass recovery, following which the sludge from both MLE systems were mixed completely to ascertain the same initial biomass concentration at the beginning of the MBR study. Once the MBR was fully operational, sludge wastage was again increased to maintain the target SRT of 20 days. During phase II (from day 100 to day 210) a membrane module was added to the aerobic zone so that the bioreactor was run as a MBR. The membrane was inserted inside the hollow cathode in the MBR–MFC while it was placed besides the open circuit cathode in the control run to demonstrate the unique filtering feature of the cathode (Fig. 1). The submerged membrane module (ZeeWeed hollow fiber membrane module, GE Water & Process Technologies, Trevose, PA, USA) was made of polyvinylidene fluoride with a nominal pore size of 0.1 lm, a total effective filtration area of 0.047 m2, and a designed membrane flux of 15.0 L m2 h1. The permeate
pumps operating on variable drives were adjusted regularly to maintain a constant flux of 4.8 L m2 h1. An air pump supplied compressed air to the built-in orifices at the bottom of each membrane module at a constant air flow rate of 6 L min1 to support aeration and control membrane fouling. Throughout the study period, magnetic stirrers were used to provide additional mixing of the mixed liquor in the anoxic and oxic chambers. 2.2. Sludge activity measurements Activated sludge activities were assessed with the specific oxygen uptake rate (SOUR) measurements in duplicate samples. Briefly, aliquots of sludge collected from the aerobic chamber in each bioreactor were aerated with pure oxygen and transferred to the respirometric bottles, which were tightly capped after insertion of a DO probe in each bottle with no headspace (air). At a predetermined time between 500 and 600 s, an aliquot of substrate (10 mg L1 NH+4–N or 20 mg COD L1 of acetate) was injected with a 10 lL glass syringe. The decrease in DO concentration in the respirometric bottle due to substrate oxidation was measured by a DO monitor (YSI model 5300A, Yellow Springs, OH, USA) and continuously recorded at 4 Hz by LabView software interfaced with a computer. The SOUR was calculated by linear regression analysis. 2.3. Membrane flux monitoring and fouling control The trans-membrane pressure (TMP) and permeate flux of the membrane module in each MBR (i.e., MBR–MFC and the open circuit control) were closed monitored using an inline battery-powered digital pressure gauge (model EW-68935-18, Cole-Parmer, Vernon Hills, IL, USA) during the phase II study. The membrane flux was maintained relatively constant at 4.8 L m2 h1 by adjusting the speed of the permeate pump. When the TMP increased and
154
S. Gajaraj, Z. Hu / Chemosphere 117 (2014) 151–157
exceeded 45 kPa (maximum allowable value of 50 kPa as suggested by the manufacturer), the membrane module was taken out of the MBR for physical cleaning. The membrane module was rinsed with approximately 30 L of distilled water under pressure to remove the attached biofilm. The membrane module was then soaked in distilled water for 0.5 h before it was submerged back to the bioreactor. No chemical cleaning was performed. 2.4. Analytical methods and statistical analysis Influent and effluent water quality was monitored twice a week during the study period. The wastewater effluent (phase I) and the permeate water (phase II) parameters such as NH+4–N, NO 3 –N, NO 2 –N, and COD were analyzed according to the standard methods (APHA, 1998). The biomass concentration, sludge volume index (SVI) in the MLE systems, and MBRs were measured in duplicate (APHA, 1998). Voltage generated by the MFC assisted system was monitored in real-time at 1 Hz using LabView data acquisition system (National Instruments, TX, USA) and the voltage was averaged every 24 h. Coulombic efficiency (gc) defined as the ratio of electrons used as current to the maximum electron production (Logan et al., 2006) was calculated as follows:
gc ð%Þ ¼
MI 8I ¼ FbqDCOD FqDCOD
where M = 32 (mol. wt. of oxygen), I is the current (A), F is Faradays constant, b = 4 (the number of electrons exchanged per mol of oxygen), q is the flow rate (L s1), and DCOD is the difference between the influent and effluent COD. The analysis of variance (ANOVA) was applied to compare the means of different groups of the study where p values less than 0.05 were considered statistically significant. 2.5. Nitrifying bacterial community analysis Terminal restriction fragment length polymorphism (T-RFLP) was used to analyze and compare the effect of MFC assisted wastewater treatment systems on the nitrifying community structure (nitrification potential) between the MLE–MFC in phase I and MBR–MFC in phase II and the open circuit control, based on the known 16S rRNA genes of ammonia-oxidizing bacteria (AOB) as described previously (Siripong and Rittmann, 2007). Activated sludge samples were collected from the aerobic chamber in the MLE and MBR systems on day 80 and day 160 respectively. Detailed information about DNA extraction and T-RFLP analysis is described in Supporting Information.
Fig. 2. Voltage output from the MLE–MFC (phase I) and MBR–MFC (phase II) over the study period.
NH+4–N. Furthermore, no nitrites were detected in the effluent. The effluent NO 3 –N concentrations in the MLE–MFC (phase I) and MBR–MFC system (phase II) were 4.7 ± 1.3 and 6.2 ± 1.5 mg L1, respectively. By contrast, the NO 3 –N concentrations in the open circuit controls were 6.9 ± 1.8 and 7.8 ± 1.5 mg L1, respectively (Fig. 3). There was no significant difference in effluent nitrate concentration between the regular MLE and MBR processes (p = 0.22). However, regardless of the type of process used in wastewater treatment, the effluent NO 3 –N concentrations of the MLE–MFC and MBR–MFC were significantly lower than those of the controls (p values
