Combined Electrochemical and Hypochlorite Pretreatment for

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Combined Electrochemical and Hypochlorite Pretreatment for Improving Solubilization and Anaerobic Digestion of WasteActivated Sludge: Effect of Hypochlorite Dosage Caihong Ye, Haiping Yuan,* Ziyang Lou, and Nanwen Zhu School of Environmental Science and Engineering, Shanghai Jiao Tong University, Shanghai 200240, People’s Republic of China ABSTRACT: The electrochemical (EC) combined with hypochlorite (ClO−) pretreatment has been proven to be effective in improving disintegration of waste-activated sludge (WAS) and enhancing anaerobic digestion recently. The ClO− dosage had a significant impact on the performance of WAS solubilization and biogas production in terms of the disintegration degree (DDCOD), release of proteins (PNs) and polysaccharides (PSs), soluble nitrogen and phosphate, and subsequent anaerobic digestion. The results showed that a higher ClO− dosage was favorable to WAS solubilization and EC−ClO− pretreatment was able to rupture the floc structure and break cells apart. According to batch anaerobic digestion tests, the optimum ClO− dosage for EC−ClO− pretreatment was 0.6% (v/v), with up to 61.1% higher methane production compared to the unpretreated sludge after mesophilic anaerobic digestion for 40 days. Model-based analysis suggested that EC−ClO− pretreatment could enhance the hydrolysis process as well as methane potential. Moreover, EC−ClO− pretreatment at the ClO− dosage of 0.6% was economically feasible based on the economic evaluation.

1. INTRODUCTION The continuous increase of sludge derived from wastewater treatment plants (WWTPs) encourages researchers and engineers to pay more attention to particular aspects of its management, especially recycling and waste-to-energy issues.1 Anaerobic digestion is considered to be the most energyefficient method for the destruction and stabilization of wasteactivated sludge (WAS).2 Moreover, the methane byproduct as a form of fuel may reduce treatment cost. However, because WAS is complex and has a rigid floc structure,3 the hydrolysis rate is often very slow and the yield of methane is limited.4,5 In this sense, disintegration is developed as the pretreatment process to accelerate the hydrolysis rate and improve the anaerobic digestion process.6 A number of pretreatment methods, such as microwave,3 ultrasonic irradiation,7 thermal,8 acidic,9 alkaline,10 and enzymatic,11 have been previously studied. After pretreated by these methods, sludge flocs were disintegrated and/or microbial cell walls were disrupted; the intracellular substances were then released into the liquid phase. The released compounds can be more readily biodegradable; thus, the methane production is promoted.4 However, physical pretreatments demand high-energy input at an industrial scale; for chemical pretreatment, a high concentration of chemicals will result in high cost and secondary pollution risk. Biological pretreatment is a very slow process and requires careful control of operating conditions.12 Considering the drawbacks of the conventional pretreatment methods mentioned above, electrochemical (EC) technology is being introduced as an alternative pretreatment strategy. Very recently, EC has shown to be effective in the membrane treatment process,13 conditioning of sludge dewaterability,14 and dewatering.15 Some researchers reported that the sludge cells can be ruptured and intracellular substances can be solubilized by EC pretreatment.16 However, some other © 2016 American Chemical Society

researchers demonstrated the limited enhancement of sludge disintegration by EC pretreatment. According to Yuan et al.,14 after EC pretreated by voltage of 15−20 V and reaction time of 15−20 min, the disintegration degree (DDCOD) of WAS was 1.3−2.0%. Gharibi et al.17 reported only 0.8−6.1% DDCOD after 20 min of detention time at 10−60 V. Therefore, more studies on EC pretreatment are still desirable. Recently, the tendency of enhancing sludge disintegration and biogas production is to combine alternatives, especially assisted by chemical methods.18 The required quantity of chemicals and energies can be reduced by combined treatments. Moreover, sludge anaerobic digestion can be sharply improved by the synergistic action, and thus, it is economically attractive for a combination of different pretreatment methods.19 A recent research performed by Zhen et al.19 verified that both WAS disintegration and anaerobic digestion could be enhanced beyond either treatment alone by the combination of electrolysis and alkaline pretreatment. Hypochlorite (ClO−) is widely used on water disinfection, fabric bleaching, surface purification, and odor removal.20 When used in water disinfection, chlorine can react with the materials of the cell wall. For example, the proteins within the cell walls may be oxidized by the reactions between chlorine and Nterminal amino groups. This will change the strength of cell walls, and thus, the cell is killed.21 In our previous studies, EC pretreatment with the aid of ClO− was compared to other pretreatment methods (thermal, alkaline, and thermal−alkaline pretreatment) and showed the superiority in sludge biogas production.22,23 However, the effect of different process conditions on EC combined with ClO− (EC−ClO−) pretreatment was not discussed. Received: December 9, 2015 Revised: March 26, 2016 Published: March 28, 2016 2990

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without any pretreatment. This unpretreated sample was also set as a control in the anaerobic digestion tests. The sludge pretreatment efficiency is always indicated by DDCOD. It can be calculated by eq 117,19

Therefore, this study is aiming to investigate the impact of different ClO− dosages on WAS solubilization and anaerobic digestion performance. First, the soluble chemical oxygen demand (COD), DDCOD, protein (PN), polysaccharide (PS), soluble nitrogen, and soluble phosphorus in pretreated and unpretreated sludges were compared to assess the efficiency of pretreatment. Second, batch tests were carried out to investigate the effect of ClO− dosage on the performance of anaerobic digestion. Finally, the variations of methane production for different ClO− dosages were interpreted using model-based analysis.

DDCOD (%) =

SCODP − SCOD0 × 100 TCOD0 − SCOD0

(1)

where SCODP represents the soluble chemical oxygen demand of pretreated sludge, SCOD0 represents the soluble chemical oxygen demand of unpretreated sludge, and TCOD0 represents the total chemical oxygen demand of unpretreated sludge. 2.3. Batch Anaerobic Digestion Tests. This assay was conducted in mesophilic temperature conditions to assess the anaerobic digestion performance of pretreated sludges. A series of 2 L conical flasks was used for the experiments. The reactors were initially filled with 1.5 L of inoculum and pretreated sludge. The volume ratio of seed sludge and pretreated sludge was 1:1. The control reactor was filled with the sludge without any pretreatment. To achieve the methane production of seed sludge, a mixture of 750 mL of milli-Q water and 750 mL of inoculum was set as the blank. The pH of reactors was modulated to about 7.0 only at the initiation of the tests by 5 mol/L NaOH or HCl. As soon as the reactors were loaded, pure nitrogen gas was used to purge the headspace of each digester for about 3 min to expel oxygen. Then, all digesters were tightly closed with butyl rubber stoppers and solid paraffin wax. All of the reactors were incubated in an air-bath shaker (100 ± 1 rpm, ZHWY-211B, Shanghai, China) at 35 ± 1 °C. The produced biogas was collected in a 2 L Tedlar gas bag, which was attached to a gas outlet on each reactor. The digestion period was 40 days. Duplicate samples were conducted to guarantee the reliability of the results, and the results were expressed as the mean ± standard deviation. Daily biogas production was measured by a 1 L glass syringe, and then the related specific cumulative methane volume was calculated. A gas chromatograph (7890B, Agilent Technologies, Santa Clara, CA) was used to determine the biogas composition. The detector of the gas chromatograph was a thermal conductivity detector (TCD) at a temperature of 250 °C, and the column was a G3591 column at a temperature of 80 °C. The sample injection volume was 100 μL, and the injection temperature was 200 °C. Nitrogen was used as the carrier gas, and the flux was 40 mL/min. 2.4. Analyses. For the determination of supernatant characteristics, a centrifuge was first applied at 12 000 rpm for 5 min to obtain the liquid phase and then a 0.45 μm mixed cellulose ester membrane was used to filter the supernatant. The filtrate was collected and analyzed for SCOD, soluble total nitrogen (TN), ammonia nitrogen (NH4+-N), nitrate (NO3−-N), nitrite (NO2−-N), soluble total phosphate (STP), soluble PN, and soluble PS. The standard reflux titrimetric method was used to measure TCOD and SCOD.24 TN, NO3−-N, NO2−-N, NH4+-N, and TP were analyzed according to standard methods.25 Soluble PS was detected according to the anthrone−sulfuric acid method, and the standard was glucose.26 The Coomassie Brilliant Blue G-250 method was applied to measure soluble PN with bovine serum albumin as the standard.27 2.5. Batch Anaerobic Digestion Modeling. The ultimate specific methane volume (M0) and hydrolysis rate constant (k) are two important parameters bound up with anaerobic digestion. They could be used to assess and compare the anaerobic digestion potential of sludge pretreated by different ClO− dosages. They were calculated according to the first-order kinetic model,19,28 which was shown in eq 2

2. MATERIALS AND METHODS 2.1. Sludges. The sludges were gained from a secondary sedimentation tank of a municipal wastewater treatment plant (WWTP) located in Shanghai, China, which is used by an anaerobic−anoxic−oxic process. The obtained sludges were delivered to laboratory as soon as sampling and then screened out coarse materials by sieved through 0.5 mm sieves. Prior to use, the sludges were centrifuged to thicken to about 3% total solid (TS) and preserved at 4 ± 1 °C to maintain freshness. The characteristics of sludges were as follows: TS of 31.5 ± 1.2 g/L, volatile solid (VS) of 21.7 ± 1.1 g/L, soluble chemical oxygen demand (SCOD) of 603 ± 92 mg/L, and total chemical oxygen demand (TCOD) of 29 181 ± 1892 mg/L. The seed sludge (inoculum) was obtained from a lab-scale anaerobic digester, which was long-term continuously run under mesophilic (35 ± 1 °C) conditions. The features of the inoculum were TS of 34.4 ± 0.8 g/L, VS of 25.5 ± 1.3 g/L, TCOD of 40 250 ± 2460 mg/L, and SCOD of 8120 ± 810 mg/L. 2.2. Pretreatment. The pretreatment experiments were carried out at room temperature in 1000 mL glass cells, in which the dimension was 10 × 10 × 10 cm3 (Figure 1). The anode and cathode

Figure 1. Schematic diagram of the pretreatment process.

were Ti/RuO2 meshes. The dimension of electrodes was 10.0 × 13.0 cm2, and the electrode distance was 5 cm in all tests. The effective electrode areas were 40 cm2. A direct current (DC) power supplier (WYJ, 5 A, 60 V, Shanghai, China) was connected to the electrodes by copper wires. The sodium hypochlorite (NaClO) solution was obtained from Sinopharm Chemical Reagent Co., Ltd. The activate chlorine content was 0.8 mol/L. The sludge volume used for pretreatment was 400 mL, and NaClO dosage was 0.2, 0.4, 0.6, 0.8, and 1.0% [volume ratio (v/v) of NaClO/sludge]. The mixture of NaClO and sludge was first stirred by a magnetic stirrer at a constant rotation speed for 10 min to ensure homogeneity. Then, a voltage of 20 V was applied to sludge samples for 40 min. The sludge pH was adjusted to around 8.0 before electrolysis. After electrolysis, samples were agitated for 30 min and standing for 24 h before batch anaerobic digestion tests. During the standing process, all of the sludge samples were stirred for 10 min every 6 h to continue the reaction between residual hypochlorite and sludge substrate and eliminate its toxic actions on anaerobic microorganisms. To eliminate the influence of agitating to sludge solubilization, the same stirring operation was conducted to the sludge

M(t ) = M 0(1 − e−kt )

(2)

where M0 represents the methane potential of sludge (mL/g of VSadded), M(t) represents the measured specific methane volume at time t (mL/g of VSadded), k is the hydrolysis rate (day−1), and t represents the time since the setup of the batch test (days). 2991

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with the organic substances and formed chloride (Cl−). This reaction would be ceased with the accumulation of Cl−. When EC was combined, Cl− would be oxidized to chlorine (Cl2) through the reaction at the anode surface and then form HClO and ClO− by the solution-phase reactions. Thus, the oxidizing reaction between ClO− and organic substances could continue, which would guarantee a high degree of sludge disintegration. Obviously, it led to some residual ClO− in the sludge sample after electrolysis was stopped. This residual ClO− would further react with the sludge substrate and was eliminated by the standing process. 3.1.2. Soluble PN and PS Releases. In activated sludge, PNs and PSs account for 70−80% of extracellular organic carbons.33,34 As shown in Figure 4, the soluble PN and PS

3. RESULTS AND DISCUSSION 3.1. Effect of ClO− Dosage on Characteristics of WAS. 3.1.1. SCOD Release and COD Solubilization. Many studies indicated that the changes of SCOD can evaluate the effectiveness of the pretreatment method.29,30 Figure 2 shows

Figure 2. Effects of ClO− dosage on SCOD and COD solubilization.

the effect of different ClO− dosages on SCOD and the degree of sludge disintegration DDCOD. When individual EC pretreatment was applied (ClO− dosage of 0), the SCOD and DDCOD were 1089 mg/L and 2.5%, respectively, which were much lower than those of EC−ClO− pretreatment. A similar result was also reported by Gharibi et al.,17 who achieved the DDCOD of only 2.5% by electrolysis treatment for 20 min at 30 V. Emna et al.18 also reported the comparable outcome in another research. Moreover, DDCOD was 11.8% when individual ClO− pretreatment was applied at a ClO− dosage of 0.8%. In comparison to EC and ClO− pretreatment alone, the combined pretreatment played an important role in the improvement of sludge disintegration. For example, SCOD increased sharply and DDCOD jumped from 8.6% at a ClO− dosage of 0.2% to 30.2% at a ClO− dosage of 0.8% by the combined pretreatment. The significant increase in DDCOD achieved by the combination of EC and ClO− pretreatment indicated an obvious synergistic effect. A parallel outcome was also obtained by Zhen et al.19 However, with a further increase of the dosage of ClO− to 1.0%, the SCOD and DDCOD were 6487 mg/L and 30.4%, respectively, and no discernible improvements were observed. This is because part of the solubilized compounds were chemical mineralization by the added oxidizing agent, which might influence the results of the methane production in this study. The synergistic mechanism of EC and ClO− can be explained by a series of chemical reactions,31,32 as shown in Figure 3. R represents the organic substances. Added ClO− first reacted

Figure 4. Effects of ClO− dosage on soluble PNs and PSs.

increased obviously when EC−ClO− pretreatment was used. The higher ClO− dosage was favorable to the PN and PS solubilization, which agree well with the SCOD results (Figure 2). When ClO− dosage was 0.2 and 0.8%, for example, soluble PN was 131.4 and 418.4 mg/L, respectively. They were 3.2 and 10.3 times higher than that of unpretreated sludge, respectively. Similarly, the soluble PS concentrations increased to 69.6 and 183.5 mg/L, which was 2.3 and 6.1 times unpretreated sludge, respectively. The increased degree was comparable to the report by Xu et al.23 It is possible that the disruption of microbial cells may cause the liberation of organic compounds (e.g., volatile fatty acids, lipids, carbohydrates, and proteins) and the increase in COD solubilization.35 The increase of soluble PN and PS implied the release of intracellular and/or extracellular components from the cells and/or extracellular polymeric substances (EPS, containing various ratios of proteins, carbohydrates, lipids/phospholipids, nucleic acids, and humic acids). These released constituents were readily available for hydrolytic bacteria and/or acidogenic microorganisms,5 thus accelerating/bypassing the hydrolysis stage. This was beneficial to the degradation of organic matters as well as methane production in the anaerobic digestion process. During the whole pretreatment process, PN was released more than that of PS under the same pretreatment conditions, as presented in Figure 4. This result was similar to other studies, which reported the pretreatment of EC alone34 or the combination of EC and alkali pretreatment.19 A possible

Figure 3. Cyclic reactions in the process of electrochemical oxidation. 2992

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Energy & Fuels explanation is that PN is the most predominant component of the microbial cell,36 and it could be released not only from EPS but also from cell lysis during the pretreatment process. The measured data of soluble PN are the sums of these two parts. Therefore, the protein concentration was higher than carbohydrate. This also confirmed that EC−ClO− pretreatment could denature the EPS as well as break cells. As a result, the intracellular defense system was disabled, and the structure of flocs was ruptured. 3.1.3. Soluble N and P Release. Along with organic matter release, many phosphorus and nitrogen species were also released during pretreatment of WAS samples. The release of TN and TP all increased following the EC−ClO− pretreatment, and a higher ClO− dosage led to higher concentrations of soluble TN and TP (Figure 5). This corroborated the SCOD

Figure 6. Measured and simulated specific cumulative methane volume of batch anaerobic digestion (symbols are experimental measured data, and lines are the model fit).

therefore, the first week is critical.35 The methane production sharply increased in the early stages, and the tendency began to flatten after day 10. In fact, about 70−80% of the specific cumulative methane volume of 40 days of anaerobic digestion was produced by the first 10 days. It was due to the hydrolyzing compounds released during the pretreatment, and the substrates would be more available to the microorganisms.38 This could also be ascribed to the increased total volatile fatty acid (TVFA) concentrations of the pretreatment process, which ranged between 325 and 2050 mg/L COD (data not shown) for all of the pretreated sludge samples, while it was only 188 mg/L COD for the control. Over the whole batch anaerobic digestion test, the specific cumulative methane volumes of the WAS pretreated by different ClO− dosages were all higher than that of unpretreated sludge obviously. When ClO− dosage was 0.6 and 0.8%, for instance, the specific cumulative methane volume was 348 and 363 mL/g of VSadded, respectively, which was 61.1 and 68.1% higher than that of unpretreated sludge (control, 216 mL/g of VSadded). The improvement degree was comparable to our previous work22 and higher than the study by Yang et al.,38 who adopted a combination method of electroflotation and electrooxidation to the pretreated sludge prior to anaerobic digestion and obtained 31.8−44.3% higher methane production. This emphasized the potential of EC−ClO− pretreatment for WAS anaerobic methane production by the synergistic actions. Nevertheless, to further increase the ClO− dosage to 1.0%, the initial methane production rate was increased but the effect diminished on approximately day 6. The specific cumulative methane volume was 228 mL/g of VSadded at day 4, which was higher than that of all other samples (p < 0.05). However, it was only 336 mL/g of VSadded at the end of anaerobic digestion, which was comparable to that when the ClO− dosage was 0.6 and 0.8%. This because the hydrolyzing compounds and bioavailable substrates increased with the addition of ClO− at the pretreatment process. In fact, the difference of ultimate methane production between the samples of ClO− dosage of 0.6, 0.8, and 1.0% was not statistically significant (p = 0.10 > 0.05). Moreover, it did not show large variations in the methane content for the reactors, and all of the produced biogas contained 64% methane. Therefore, consid-

Figure 5. Effects of ClO− dosage on soluble N and P.

behavior given before (Figure 2). However, particular attention should be paid to the NH4+-N release tendency. As depicted in Figure 5, NH4+-N increased with an increasing ClO− dosage and peaked at 0.6% but a decreasing tendency was observed when the ClO− dosage continued to increase. Meanwhile, the concentrations of NO3−-N in the supernatant increased with the stronger pretreatment intensity. However, the amount of NO3−-N was very low all along (less than 21.4 mg/L), and NO2−-N was not detected in the supernatant during the whole pretreatment process. Therefore, organic nitrogen was supposed to be the major component of released TN. The decrease of NH4+-N can be ascribed to the reaction between hypochlorite and ammonium:37 2NH4+ + 3HOCl → N2 + 5H+ + 3Cl− + 3H2O. 3.2. Effect of ClO− Dosage on Batch Anaerobic Digestion Tests. The effect of EC−ClO− pretreatment on WAS anaerobic digestion at different ClO− dosages was researched in this assay, and the optimal dosage of ClO− was determined. As seen in Figure 6, the specific methane volume was high at the first day of anaerobic digestion, which means that none of the pretreated digesters had a lag phase. Yu et al.22 reported that hypochlorite had adverse effects on anaerobic microbial activity, and this would led to a low rate of methane production. The increased initial methane production in this study may be attributed to the standing process before anaerobic digestion. The maximum utilization of substrate happened at the first 5−7 days of a batch anaerobic digestion; 2993

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hydrolytic bacteria, and thus, the hydrolysis stage was accelerated. The values of M0 increased significantly (p = 0.003 < 0.05) with an increasing ClO− dosage, except for the case of 1.0%, where a decreased M0 was determined. A different tendency was reported by Zhen et al.,19 who revealed that the enhancement of the sludge ultimate methane yield (M0) by electrical−alkali pretreatment was very slight, probably because the sludge types and properties of the sludge were different and also as a result of the variation of the additive chemical agent. The higher COD solubilization (Figure 2) corresponding to a higher k and a lower M0 at the ClO− dosage of 1.0% indicated that part of the solubilized compounds was turned to refractory fractions or non-biodegradable components and was not converted to methane at the anaerobic digestion process. A similar behavior was also noticed by Wang et al.4 with heat pretreatment and Kim et al.40 with ultrasonics combined with alkaline pretreatment. 3.4. Economic Analysis. An economic analysis was carried out according to the experimental results achieved in this study by comparing the different pretreatment processes to the control. The results were shown in Table 3. The energy content

ering the economy of the agent, 0.6% was considered to be the optimum ClO− dosage. Table 1 illustrated the final TS, VS, and TCOD removal for the digested sludge under different pretreatment methods. Table 1. Degradation of Organic Components after Anaerobic Digestion (with 95% Confidence Intervals) control TS removal (%) VS removal (%) TCOD removal (%) soluble PN before digestion (mg/L) soluble PN after digestion (mg/L) soluble PS before digestion (mg/L) soluble PS after digestion (mg/L)

31.9 38.2 38.5 40.6

± ± ± ±

0.6 1.0 1.3 3.3

0% (v/v) 32.0 38.6 39.7 122.5

± ± ± ±

0.6% (v/v)

0.9 0.7 1.5 5.8

32.1 42.7 44.8 351.9

± ± ± ±

0.8 1.2 0.9 7.1

56.8 ± 2.8

84.7 ± 6.1

138.9 ± 5.5

29.9 ± 2.9

50.4 ± 2.1

123.7 ± 6.6

48.0 ± 7.9

68.6 ± 5.4

95.5 ± 4.3

Adding ClO− into sludge notably promoted the fermentation performance with respect to the degradation of organic solids, corresponding well with the results measured for methane production (Figure 6). The average TS, VS, and TCOD removals for the control were about 31.9 ± 0.6, 38.2 ± 1.0, and 38.5 ± 1.3%, respectively. However, all of them increased to 32.1 ± 0.8, 42.7 ± 1.2, and 43.8 ± 0.9%, respectively, when ClO− dosage was 0.6%. One reason for a higher organic solids removal achieved for EC−ClO− pretreatment might be the higher content of dissolved organic matter in the pretreated sludge compared to the control.39 As shown in Table 1, the soluble PN and PS decreased for the EC−ClO− pretreated sludge after anaerobic digestion. The results indicated that these compounds were partly degraded during the batch tests. However, soluble PN and PS for the control increased in comparison to that before anaerobic digestion. The reason was due to the production of these compounds being more than the biodegraded amount in the digestion system. Nevertheless, soluble PN and PS of the pretreated sludge were higher than that of unpretreated sludge at the end of anaerobic digestion, which implied an accumulation of soluble organic matter in the liquid phase. These need further investigation and consideration for the treatment of the filtrate. 3.3. Kinetic Model. The fitting of experimental specific cumulative methane production data to the first-order model was shown in Figure 6. The experimental data anastomose well with the model (R2 > 0.98 in all studied cases). The estimated k and M0 were shown in Table 2. According to Table 2, the hydrolysis rate constant (k) increased (p = 0.025 < 0.05) with the increase of the ClO− dosage. This is presumably owed to the release of organic matter, which was readily available for

Table 3. Evaluation of the Economy 0% (v/v)

Energy Balance (Per Ton of Dry Solid) increased biogas production 30 167 energy content of increased biogas 195.0 1085.5 energy applied in pretreatment 300.0 633.3 net energy yield (+)/applied (−) −105.0 452.2 Cost Estimation (Per Ton of Dry Solid) energy profit (+)/cost (−) −11.13 47.93 ClO− cost 0.00 24.40 net profit (+)/cost (−) −11.13 23.53

unpretreated 0 0.2 0.4 0.6 0.8 1

k (day−1) 0.18 0.19 0.19 0.20 0.22 0.22 0.28

± ± ± ± ± ± ±

0.01 0.01 0.05 0.01 0.00 0.01 0.01

± ± ± ± ± ± ±

$ $ $

4. CONCLUSION The present paper investigated the effect of EC−ClO− pretreatment on WAS solubilization and biogas production at different ClO− dosages. The results showed that a higher ClO− dosage was favorable to sludge solubilization. According to the batch anaerobic digestion tests, 0.6% (v/v) was chosen to be the optimum ClO− dosage for EC−ClO− pretreatment with up to 61.1% methane production improvement compared to that of unpretreated sludge at the end of anaerobic digestion. Continuing to increase the ClO− dosage to 0.8 and 1.0% did

M0 (mL/g of VSadded) 209 238 269 303 340 361 323

m3 kWh kWh kWh

of biogas with 65% CH4, and the cost of electricity was estimated at 65 kWh/m341 and $0.106/kWh,42 respectively. The cost of NaClO was estimated at $0.122/L according to the market price. The energy consumption in the pretreatment stage by individual EC (ClO− dosage of 0%) was significantly lower than that of EC−ClO− (ClO− dosage of 0.6%) pretreatment. However, the increased biogas production owed to the individual EC pretreatment could not offer enough energy to keep the EC device. For EC−ClO − pretreatment, although it required considerably high energy, the increment in biogas production was capable of covering the extra costs. It was seen that EC−ClO− pretreatment at the ClO− dosage of 0.6% is economically feasible, with net savings of $23.53/ton of dry solid. Although the variation of sludge dewaterability and the decrease of disposed TS would impact the estimated cost, the results of this brief calculation showed the attraction of EC−ClO− pretreatment.

Table 2. Estimated k and M0 at Different ClO− Dosages by the First-Order Model (with 95% Confidence Intervals) ClO− dosage (%, v/v)

0.6% (v/v)

1 2 2 2 2 3 2 2994

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(13) Ibeid, S.; Elektorowicz, M.; Oleszkiewicz, J. A. Modification of activated sludge properties caused by application of continuous and intermittent current. Water Res. 2013, 47, 903−910. (14) Yuan, H. P.; Yan, X. F.; Yang, C. F.; Zhu, N. W. Enhancement of waste activated sludge dewaterability by electro-chemical pretreatment. J. Hazard. Mater. 2011, 187, 82−88. (15) Mahmoud, A.; Olivier, J.; Vaxelaire, J.; Hoadley, A. F. Electrical field: a historical review of its application and contributions in wastewater sludge dewatering. Water Res. 2010, 44, 2381−2407. (16) Song, L. J.; Zhu, N. W.; Yuan, H. P.; Hong, Y.; Ding, J. Enhancement of waste activated sludge aerobic digestion by electrochemical pre-treatment. Water Res. 2010, 44, 4371−4378. (17) Gharibi, H.; Sowlat, M. H.; Mahvi, A. H.; Keshavarz, M.; Safari, M. H.; Lotfi, S.; Bahram Abadi, M.; Alijanzadeh, A. Performance evaluation of a bipolar electrolysis/electrocoagulation (EL/EC) reactor to enhance the sludge dewaterability. Chemosphere 2013, 90, 1487−1494. (18) Feki, E.; Khoufi, S.; Loukil, S.; Sayadi, S. Improvement of anaerobic digestion of waste-activated sludge by using H2O2 oxidation, electrolysis, electro-oxidation and thermo-alkaline pretreatments. Environ. Sci. Pollut. Res. 2015, 22, 14717−14726. (19) Zhen, G.; Lu, X.; Li, Y.-Y.; Zhao, Y. Combined electrical-alkali pretreatment to increase the anaerobic hydrolysis rate of waste activated sludge during anaerobic digestion. Appl. Energy 2014, 128, 93−102. (20) Asokan, K.; Subramanian, K. Design of a tank electrolyser for insitu generation of NaClO. Proceedings of the World Congress on Engineering and Computer Science 2009 (WCECS 2009); San Francisco, CA, Oct 20−22, 2009; Vol. I. (21) AK, S.; C, V. Effect of prechlorination on coagulation of algae. Indian J. Environ. Health. 1979, 21, 16−22. (22) Yu, B.; Xu, J.; Yuan, H.; Lou, Z.; Lin, J.; Zhu, N. Enhancement of anaerobic digestion of waste activated sludge by electrochemical pretreatment. Fuel 2014, 130, 279−285. (23) Xu, J.; Yuan, H.; Lin, J.; Yuan, w. Evaluation of thermal, thermalalkaline, alkaline and electrochemical pretreatments on sludge to enhance anaerobic biogas production. J. Taiwan Inst. Chem. Eng. 2014, 45, 2531−2536. (24) State Environmental Protection Administration of China. Analytical and Monitoring Methods of Water and Wastewater, 4th ed.; China Environmental Science Press: Beijing, China, 2002. (25) APHA, AWWA, and WEF. Standard Methods for the Examination of Water and Wastewater, 21st ed.; American Public Health Association (APHA), American Water Work Association (AWWA), and Water Environment Federation (WEF): Washington, D.C., 2005. (26) Koehler, L. H. Differentiation of Carbohydrates by Anthrone Reaction Rate and Color Intensity. Anal. Chem. 1952, 24, 1576−1579. (27) Bradford, M. M. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 1976, 72, 248−254. (28) Zhou, X.; Wang, Q.; Jiang, G. Enhancing methane production from waste activated sludge using a novel indigenous iron activated peroxidation pre-treatment process. Bioresour. Technol. 2015, 182, 267−271. (29) Apul, O. G.; Sanin, F. D. Ultrasonic pretreatment and subsequent anaerobic digestion under different operational conditions. Bioresour. Technol. 2010, 101, 8984−8992. (30) Chen, Y.; Jiang, S.; Yuan, H.; Zhou, Q.; Gu, G. Hydrolysis and acidification of waste activated sludge at different pHs. Water Res. 2007, 41, 683−689. (31) Scialdone, O.; Randazzo, S.; Galia, A.; Silvestri, G. Electrochemical oxidation of organics in water: role of operative parameters in the absence and in the presence of NaCl. Water Res. 2009, 43, 2260− 2272. (32) Szpyrkowicz, L.; Radaelli, M.; Daniele, S. Electrocatalysis of chlorine evolution on different materials and its influence on the performance of an electrochemical reactor for indirect oxidation of pollutants. Catal. Today 2005, 100, 425−429.

not achieve significant improvements on ultimate methane production, probably as a result of the composition change of soluble compounds. Model-based analysis implied that both the hydrolysis rate and biogas production were improved by EC− ClO− pretreatment. On the basis of the economic evaluation relative to conventional anaerobic digestion, EC−ClO − pretreatment at the ClO− dosage of 0.6% is economically feasible and could achieve $23.53/ton of dry solid of net savings.

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This work was supported by the National Natural Science Foundation of China (51208295 and 51178261), the Shanghai Science and Technology Commission (12231202101), and the Ministry of Education Institution of Higher Learning Doctor Discipline End Scientific Research Fund (20120073120050).

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DOI: 10.1021/acs.energyfuels.5b02884 Energy Fuels 2016, 30, 2990−2996

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DOI: 10.1021/acs.energyfuels.5b02884 Energy Fuels 2016, 30, 2990−2996