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The regulation and control strategies of a sequencing batch reactor for simultaneous nitrification and denitrification at different temperatures
Bioresource Technology 133 (2013) 59–67
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The regulation and control strategies of a sequencing batch reactor for simultaneous nitrification and denitrification at different temperatures Jingbo Guo a,⇑, Lanhe Zhang b, Wei Chen a, Fang Ma c, Honglei Liu b, Yu Tian b a
School of Civil and Architecture Engineering, Northeast Dianli University, Jilin 132012, China School of Chemical Engineering, Northeast Dianli University, Jilin 132012, China c State Key Lab of Urban Water Resource and Environment, School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China b
h i g h l i g h t s " Prolonged aeration favored the SBR–SND system under low temperatures. " Further extension of aeration was insubstantial, only increased operational cost. " C/N ratio adjustment to offset the temperature impacts was ineffective. " Long SRT lessened low temperature effect but reduced N removal at high temperature. " ORP, DO, SVI and EPS changed with temperature and led to performance variation.
a r t i c l e
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Article history: Received 23 November 2012 Received in revised form 5 January 2013 Accepted 5 January 2013 Available online 22 January 2013 Keywords: Simultaneous nitrification and denitrification Sequencing batch reactor Temperature Regulation and control strategies
a b s t r a c t The performance of a sequencing batch reactor for simultaneous nitrification and denitrification (SBR– SND) was investigated under 5–30 °C and strategies against temperature influences were proposed. Aeration of 8, 7, 7 and 6 h were sufficient for 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C, respectively. Further extension was insubstantial, only increased the aeration cost. The adjustment of C/N ratio to offset the temperature impacts was not remarkable. Prolonged sludge retention time lessened the influences of low temperature but deteriorated the system at high temperature. The oxidation reduction potential, the dissolved oxygen concentration, the sludge volume index and the extracellular polymeric substances amount changed with temperature alterations and thus affected the system performance. In conclusion, measures should be taken for temperature oscillations and the regulation and control of the operational parameters could alleviate the influences of temperature on the performances of the SBR–SND system. Ó 2013 Elsevier Ltd. All rights reserved.
1. Introduction The adverse environmental impacts of nitrogenous compounds include promotion of eutrophication, toxicity to aquatic organisms and depletion of dissolved oxygen (DO) in receiving water bodies due to bacterial oxidation of ammonia to nitrate (He et al., 2009). Nitrogen removal to an appropriate concentration is important and compulsory before wastewaters are being discharged into the environment (Seifi and Fazaelipoor, 2012). Biological treatment is one of the most economical processes for nitrogen removal. Biological nitrogen removal (BNR) from wastewaters is usually accomplished through sequential nitrification and denitrification processes, i.e., conventional nitrification–denitrification process, by which nitrification and denitrification are performed in separate compartments (Lan et al., 2011). In the nitrification compartment, ⇑ Corresponding author. Tel./fax: +86 432 62834526. E-mail address: [email protected] (J. Guo). 0960-8524/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2013.01.026
aerobic autotrophic nitrifiers oxidize NHþ 4 —N to NO2 —N and then NO3 -N; in the denitrification compartment, NO3 —N is reduced to NO 2 —N and then to gaseous nitrogen compounds by heterotrophic denitrifying bacteria under anoxic conditions (Wan et al., 2009). Studies showed that nitrification and denitrification could occur concurrently in a single reactor under aerobic conditions, which is often referred to as the simultaneous nitrification and denitrification (SND) process (Yang and Yang, 2011). The SND process represents a significant advantage over the conventional separated nitrification and denitrification processes. First, SND system eliminates the serial operation of two separate tanks and thus requires a smaller footprint and simpler operating procedures (Yang and Yang, 2011). Favorable nitrogen removal efficiency could be achieved at lower C/N ratios as more carbon substrates are available for denitrifers in the SND system (Hocaoglu et al., 2011a). It is estimated that the SND process utilizes 22–40% less carbon source and reduces sludge yield by 30% when compared with the conventional BNR system (Seifi and Fazaelipoor, 2012). The SND
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could be accomplished under neutral pH level with less demand for alkalinity because alkalinity is consumed during nitrification but produced during denitrification (Rong et al., 2007). Also, oxygen consumption may be greatly reduced because NO 3 -N can be utilized to the fullest extent as the alternate electron acceptor and thus low DO concentration (0.10–1.00 mg/L) is sufficient for SND reaction (Chiu et al., 2007; Hocaoglu et al., 2011a,b). Therefore, to take full advantages of this process, factors including carbon and ammonium ratio (C/N) (Chiu et al., 2007), sludge retention time (SRT) (Wu et al., 2011), DO concentration (Hocaoglu et al., 2011a) and temperature (Ilies and Mavinic, 2001) as well as other factors (He et al., 2009) on the performances of the SND process had been widely investigated. As an important environmental factor of the biological wastewater treatment process, temperature not only influences the metabolic activities of the microbial population, but also has a profound effect on such factors as gas-transfer rates and the settling characteristics of the biological solids (Metcalf and Eddy Inc., 2003). Biological reaction rates increase with rising temperature until an optimal temperature is reached; above the optimal temperature, enzymatic proteins denature and the rates decrease. The nitrification and denitrification of a high ammonia landfill leachate were greatly inhibited at 10 °C (Ilies and Mavinic, 2001). Temperature and NHþ 4 -N removal rate were positive correlated with temperature in an interval of 5–20 °C (Willers et al., 1993). The nitrification rate was an increasing function of temperature in the range of 15–25 °C (Antoniou et al., 1990). Maximum nitrogen removal of a shrimp aquaculture wastewater with a SBR was obtained in the presence of 2.8–4% salinity and in a temperature range of 22–37 °C (Fontenot et al., 2007). The temperatures of municipal wastewater vary in the range of 10–30 °C and may fall below 10 °C in cold areas, for example, the North China. To date, there are limited reports about the influences of temperature on the performance of the SND system and the corresponding coping measures. In the present study, the performance of a sequencing batch reactor based simultaneous nitrification and denitrification (SBR–SND) system was investigated under different temperatures feeding with synthetic wastewater. The changes of oxidation reduction potential (ORP), DO concentration, the sludge volume index (SVI) value and the amount of extracellular polymeric substances (EPS) under different temperatures were detected. The regulation and control strategies were proposed to achieve relatively high nitrogen removal efficiency, low costs and easy maintenance. The objective of this study is to investigate the influences of temperature on the performance of the SBR–SND system and provide useful guidance for its operation in accordance with temperature oscillations, rather than following the stereotype without any alternation.
2. Methods 2.1. The lab-scale SBR and the operating conditions A lab-scale SBR with dimensions of 30 10 cm (H D) and a working volume of 2.5 L was fabricated with polymethyl methacrylate. Samples were collected from the three sampling outlets located on the SBR side. Sludge discharging outlets was located at the bottom of the SBR. A complete mixing inside the SBR was ensured by suspending the mixed liquor with a mechanical stirrer at a rate of 100 rpm. Aeration was provided by pressurized air passing through long stone diffusers and the initial DO concentration was controlled at 0.5–1.0 mg/L by air flow valve. Incubator was used to provide different temperatures. Sludge collected from the aerobic pond of a municipal wastewater treatment plant (Jilin, China) was used as inoculum. The sludge was elutriated with clean water
three times to remove fine suspended solids that may interfere with the analyses. Then the sludge was added to SBR at a final concentration of approximately 1500 mg/L in the synthetic wastewater. One typical cycle of the SBR contained four procedures: feed (instantaneous), reaction (7.5 h), settling (0.5 h), decant (instantaneous, 80% water exchange, 2 L), idle (4 h). 2.2. Synthetic wastewater Synthetic wastewater was used for investigating the SND performance at different temperatures. The composition was as follows (per liter): CH3COONa, 0.6–1.5 g; KH2PO3, 20 mg; MgSO4, 60 mg; CaCl2, 360 mg; FeSO4, 3 mg; NH4Cl, 35–50 mg; trace element solution, 0.5 mL. The trace element solution contained the following chemicals (per liter): FeCl36H2O, 1.5 g; H3BO3, 0.15 g; CuSO45H2O, 0.03 g; KI, 1.18 g; MnCl24H2O, 0.12 g; ZnSO47H2O, 0.12 g; CoCl2, 0.15 g. The characteristics of the synthetic wastewater are summarized in Table 1. 2.3. Experimental plan During the start-up period, the SBR was operated with a hydraulic retention time (HRT) of 10 h and a sludge retention time (SRT) of 20 d at 25 ± 2 °C. The influent COD concentration was approximately 450 mg/L and the TN concentration was gradually increased from 10 to 50 mg/L. After the steady-state was achieved as demonstrated by efficient and stable COD and TN removal efficiencies, the SBR was operated for an additional 20 d at 25 ± 2 °C. Then the temperature of the incubator where the SBR placed was set at 5 ± 2 °C. The effluent quality was monitored until the steady-state was achieved. Then an average value of effluent quality from seven serial cycles was presented to reveal the necessary aeration time at 5 ± 2 °C. The performances of the SBR–SND system under different C/N ratios at the same temperature were then investigated applying the necessary aeration time. Then the SRT was controlled through direct removal of sludge from the bioreactor (1/5, 1/10, 1/15 and 1/20 of the bioreactor volume) on daily basis, to maintain SRT of 5, 10, 15 and 20 d respectively for exploring the influences of SRT on the SBR–SND system. After that, the influences of the pH value were investigated by adjusting the influent pH in a range of 4.8–9.7. The performances of the SBR–SND system and the corresponding SVI and EPS amount was detected during 30 d under 5 ± 2 °C when the SND–SBR system were operated under the optimal parameters. Meanwhile, the variation of DO and ORP value in a typical cycle were detected. Then the temperature of the incubator was successively set at 10 ± 2, 20 ± 2, 30 ± 2 °C and the same experimental procedures were conducted under each temperature. Two turnovers of the SRT (20 d) were operated to reach the steady-state for each temperature change. Finally, data obtained from different temperatures were collected and compared to reveal the characteristics of the SBR–SND system under various temperatures and the corresponding regulation and control strategies. 2.4. Analytic methods The concentration of COD, NHþ 4 —N, NO2 —N, NO3 —N, TN and mixed liquor suspended solids (MLSS) were measured according
Table 1 The characteristics of the synthetic wastewater. Parameters
COD
TN
NHþ 4 -N
NO 3 -N
NO 2 -N
Range (mg/L)
100–500
40–42
40–42
0–0.5
0–0.5
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to Standard Methods for the Examination of Water and Wastewater (APHA, AWWA, WPCP, 2005). Temperature, pH, ORP and DO were analyzed online with a WTW pH/ORP sensor (WTW, pH 3310) and a DO sensor (WTW, OXi340i). The extraction and quantification of EPS were performed according to the protocol described by Liang et al. (2010). 3. Results and discussions 3.1. Start-up of the SBR The reactors became stable after being acclimated for 30 d. As shown in Fig. 1, the COD and TN removal efficiencies were up to 90% and 85%, respectively. The effluent TN concentration was 3– 8.56 mg/L when the influent TN concentration varied from 10 mg/L to 53.5 mg/L. SND was accomplished as the nitrification and denitrification occurred simultaneously in a single reactor under identical conditions. 3.2. Adjustment of reaction time under different temperatures The aeration duration was set as 10 h and the effluent COD, TN and NHþ 4 -N concentration was detected at hourly intervals when the influent NHþ 4 -N concentration was 40–42 mg/L. The data presented in Fig. 2 was an average value of seven serial cycles at the steady state of each temperature. The removal of COD could be divided into three phase. At 5 ± 2 and 10 ± 2 °C, more than 50% COD was removed within the initial 4 h; the COD removal efficiency reduced dramatically during the following 2 h; and then the effluent COD concentration hardly changed thereafter. At 20 ± 2 and
100
600 500
80
400 60
Inf. COD Eff. COD COD removal efficiency
300
40
200 20
100 0
0
5
10
15
20
25
30
COD removal efficiency (%)
COD concentration (mg/L)
(a)
0
Time (d) 60
100
50
80
40
Inf. TN Eff. TN TN removal efficiency
30
60
40 20 20
10 0
0
5
10
15
20
25
30
0
Time (d) Fig. 1. The removal of COD (a) and TN (b) during start-up period.
TN removal efficiency (%)
TN concentration (mg/L)
(b)
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30 ± 2 °C, up to 80% COD was removed within the initial 3 h; the removal rate decreased gradually in the following 2 h; little changes of COD concentration were observed thereafter. The effluent NHþ 4 -N and TN stabilized after 8, 7, 7 and 6 h operation when the respective temperatures were 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. Thus, the nitrogen removal efficiency increased with the elevation of temperature and the prolonged reaction time was beneficial for the SBR operated under low temperatures. Microorganisms could store carbon substrates and then served as carbon sources for denitrification and heterotrophic nitrification (Ding et al., 2011). This could explain the higher decreasing rate of COD at the initial stage than the nitrogen compounds and then the COD concentration stabilized, while the TN and NHþ 4 -N concentrations deceased at higher rates. There was a discrepancy between the effluent NHþ 4 -N and TN concentration, suggesting the accumulation of NO 3 -N and/or NO 2 -N. At 5 ± 2 °C, this discrepancy was small as the nitrification and the denitrification were greatly inhibited by low temperature and the majority of the added nitrogen compound was discharged without decomposition. This discrepancy decreased with the elevation of temperature from 10 ± 2 to 20 ± 2 °C and then to 30 ± 2 °C as the equilibrium of the nitrification and denitrification rates was improved. The effect of temperature on NO 3 -N and/or NO -N accumulation was probably relative to the growth charac2 teristics of nitrifiers and denitrifiers, namely, denitrifiers were more sensitive than nitrifiers to low temperature. This was in accordance with the previous study that the denitrification was a key step to regulate the overall nutrient removal efficiencies at low temperatures (Choi et al., 1998). S According to Michaelis–Menten equation V ¼ KVmmþS and ArrheEa nius equation V m ¼ AexpRT , high temperature will improve the biodegradation rate as a result of the enhanced microbial enzyme activity (Johnson et al., 2010). An increment of 10 °C from 22.5 to 32.5 °C resulted in an improvement of the ammonium oxidation rate (AOR) and nitrite oxidation rate (NOR) by a factor of 1.36 and 1.26, respectively; the AOR and NOR decreased to 44% and 55%, of the initial rates at decreasing temperatures from 22.5 to 12.5 °C (Sudarno et al., 2011). As a counteraction of the low temperature wastewater treatment, the biodegradation duration was frequently extended by prolonging hydraulic retention time (HRT) or aeration time (Kayranli and Ugurlu, 2011). Chiemchaisri and Yamamoto (1993) found that by increasing aeration time in operational cycle of a membrane separation bioreactor for on-site domestic wastewater treatment, the nitrification could be recovered at low temperatures as a result of the enhanced oxygen supply. Low temperature not only influences the metabolic activities of nitrifiers and denitrfiers, but also affects the bacterial community and the species richness. Karkman et al. (2011) identified the bacteria in a bioreactor treating inorganic mine waters under 5 and 10 °C and found that the dominant nitrifiers were distinct at different temperature; meanwhile, as the temperature increased from 5 to 10 °C, the operational taxonomic units (OTUs) elevated from lower than 150 to higher than 300, demonstrating that the bacterial species richness enhanced with the elevation of temperature. In the present case, 8, 7, 7 and 6 h were sufficient when the respective temperatures were 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. The further extension of reaction time was insubstantial for performance improvement, especially at low temperature conditions, only increased the operational cost for aeration. Besides, shorter reaction time could provide more degradable organic substrates as reducing power for denitrification in a SND system; otherwise, the readily-biodegradable COD was oxidized by oxygen in the bioreactor (Liu et al., 2010). Thus, for optimizing SND efficiency and reducing operational cost, an appropriate reaction time should be determined under different temperature conditions and the pro-
J. Guo et al. / Bioresource Technology 133 (2013) 59–67
50
300 30 200
20
100
+
10
COD concentration (mg/L)
NH4 -N and TN concentration (mg/L)
Eff. TN Eff. COD
40
400
+
5±2 oC 0
NH4 -N and TN concentration (mg/L)
+
Eff. NH4 -N
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Eff. NH4 -N Eff. TN Eff. COD
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Time (d)
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COD concentration (mg/L)
400
NH4 -N and TN concentration (mg/L)
+
NH4 -N and TN concentration (mg/L)
Eff. TN Eff. COD
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50
+
Eff. NH4 -N
0
4
10
0
Time (d)
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400
COD concentration (mg/L)
50
+
Eff. NH4 -N Eff. TN Eff. COD
40
400
300 30 200
20
100
10
COD concentration (mg/L)
62
30±2 oC 0
0
2
4
6
8
10
0
Time (d)
Fig. 2. The time courses of effluent NHþ 4 , TN and COD concentrations under different temperatures.
longed reaction time was not a radical solution for low temperature wastewater treatment. 3.3. Adjustment of influent C/N ratio under different temperatures In SND system, less carbon was demanded as nitrification and denitrification was accomplished in a single reactor when compared with the conventional BNR processes. As the autotrophic nitrification is generally slower than the heterotrophic metabolism, SND requires a degradable organic substrate to provide reducing power for heterotrophic nitrification and denitrification processes (Liu et al., 2010). Nevertheless, competition between autotrophic and heterotrophic microorganisms would be induced at high organic loading rate (OLR). In a sequencing batch airlift reactor, the competition for oxygen between heterotrophic and autotrophic bacteria became critical at OLR higher than 2 kg COD/(m3 d) (Wan et al., 2009). Hence, the C/N ratio should be adjusted to achieve the equilibrium between nitrification and denitrification processes. As shown in Fig. 3, there was an optimal C/N ratio for the removal of organics and the nitrogen compounds at different temperatures. The highest COD, NHþ 4 -N and TN removal efficiencies at 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C were obtained when the influent C/N ratios were 8, 8, 10, 10, respectively, indicating that C/N ratio requirements varied at different temperatures. Low C/N ratio resulted in a rapid carbon deficit, causing an unbalanced SND process in SBR. Complete removal of NHþ 4 -N was achieved when the initial C/N was adjusted to 11.1 at 25 ± 2 °C (Chiu et al., 2007). On the other side, in a compact suspended carrier biofilm reactor (SCBR) developed for SND, the diversity of microbial community structure and the population of nitrifiers was in inverse propor-
tions to C/N ratio (Xia et al., 2008). Therefore, it’s necessary to choose an appropriate C/N ratio according to the types of SND process being adopted and wastewater being treated. However, the nitrogen removal efficiencies obtained at various influent C/N ratios of a certain temperature was not remarkable, suggesting that the adjustment of C/N ratio to offset the impacts of temperature was not an effective option. 3.4. Adjustment of sludge retention time (SRT) under different temperatures As shown in Fig. 4, the best performances were obtained with a 20, 15, 10 and 10 d SRT at 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. The removal efficiency of NHþ 4 -N was higher than that of TN at all condi tions, indicating the accumulation of NO 3 -N and/or NO2 -N. The biomass concentration in the biological reactor is a function of the SRT and the different biomass concentrations resulted from the variations of SRT was crucial for the mass transfer conditions between DO, nitrogen compounds and biomass fractions responsible for nitrifications and denitrification (Hocaoglu et al., 2011b). Long SRT could reinforce nitrification as a result of biomass enhancement and better retention of slow growing microorganisms. In a membrane bioreactor for the treatment of synthetic municipal wastewater, the highest total nitrogen removal and nitrogen removal via SND were achieved with a SRT of 40 d under 16 ± 1 °C and a DO concentration of 0.5–1.5 mg/L (Holakoo et al., 2007) and decreased with a SRT of 20 d. Holakoo et al. (2007) indicated the deteriorated SND with decreased SRT primarily attributed to a reduction of the floc size, which limited denitrification as the overall anoxic volume fraction of the flocs reduced. SRT extension was one of the common measures to sustain the biolog-
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100
100 80
TN
60 40 20 0
C/N=10
C/N=8
C/N=5
0
COD + NH4 -N
5±2 oC
Removal efficiency (%)
Removal efficiency (%)
COD + NH4 -N
7
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TN
60 40 20
C/N=12
21
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28
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COD + NH4-N
40
TN C/N=10
C/N=8
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14
21
Removal efficiency (%)
(%)
100
Removal efficiency
C/N=12
Time (d)
Time (d)
0
C/N=10
C/N=8
C/N=5
80
TN
30±2 oC
40 C/N=5
C/N=12
C/N=10
C/N=8
20 0
28
COD + NH4-N
60
0
7
14
21
28
Time (d)
Time (d)
Fig. 3. Removal of COD, NHþ 4 and TN under different effluent C/N ratios.
100
100 COD + NH4 -N
5±2 C
80
TN
Removal efficiency (%)
Removal efficiency (%)
80
60
40
20
0
5
COD + NH4 -N
o
10
15
TN
60
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20
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SRT (h) 100
COD + NH4 -N
40
20
5
20
30±2 oC
TN
Removal efficiency (%)
Removal efficiency (%)
COD + NH4 -N
100
20±2 oC
60
0
15
SRT (h)
TN
80
10±2 oC
10
15
SRT (h)
20
80
60
40
20
0
5
10
15
SRT (h)
Fig. 4. Removal of COD, NHþ 4 and TN under different SRTs.
20
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ical wastewater treatment performance and stability under low temperatures as increased sludge age could lessen the unfavorable effect of low temperatures (under 15 °C) and stabilize the nitrification process (Komorowska-Kaufman et al., 2006). However, as shown in Fig. 4, the carbon and nitrogen removal efficiencies were not always proportional to the SRT as inflection points were detected at SRT of 15 or 10 d when temperatures were higher than 5 ± 2 °C. Long SRT would accelerate the floc breakup, especially under high temperature conditions (Holakoo et al., 2007). Thus, the extension of SRT in a suitable range could lessen the influences of low temperature but the biological nitrogen capability of the SBR–SND system would be detrimentally affected by longer SRT at higher temperatures. 3.5. Adjustment of pH under different temperatures In the conventional nitrification and denitrification system, the generation of acid compounds during the nitrification process will cause pH reduction; on the contrary, the pH increases in the denitrification process as a result of the alkali formation. The pH was often taken as an indicator of the progressing degree of nitrification and denitrification. In the SND system, the alkalinity generated by the denitrification could partially compensate for that consumed by the nitrification process. As shown in Fig. 5, the best performance of the SBR was obtained with pH 8 at all temperatures, suggesting that neutral to alkali conditions were beneficial for the conversion of organics and the nitrification and denitrification of the nitrogen compounds. This was consistent with the results obtained by He et al. (2009) that the optimum TN removal was detected at around pH 7.2 with an investigated range from 4.8 to 9.7.
3.6. Variation of DO concentration and ORP under different temperatures The saturated DO concentration in water increases with the reduction of temperature and thus the DO concentration in the SBR will vary with temperature fluctuations. The relationship between the saturated DO concentration and the DO concentration detected in the reactor could be represented by the equation of oxygen transfer rate: ddDOt ¼ K La ðDOs DOÞ. The KLa value increases with the elevation of temperature. On the contrary, the saturated DO concentration decreases as the temperature enhances. Therefore, the temperature has reverse effects on the oxygen transfer rate. However, these effects could not be offset and generally the low temperature is beneficial for oxygen transfer. As shown in Fig. 6, the DO concentration showed a slight decrease at the initial stage, while longer durations of the decreasing trend were observed at higher temperatures. Holman and Wareham (2005) demonstrated that the depletion of COD and NHþ 4 -N caused a sudden DO increase, which was also observed in the present study under 20 ± 2 and 30 ± 2 °C. The increase of DO concentration at 5 ± 2 and 15 ± 2 °C emerged earlier than at higher temperatures, which were the joint results of higher saturated DO concentration and lower metabolic activities of microorganisms under low temperatures. A decrease in microbial activity would cause a decrease in oxygen and substrate utilization and resulted in high DO and effluent COD and nitrogen concentration at low temperature conditions. The increase in DO would, in turn, hamper denitrification, which requires anaerobic environments in order to convert NO 3 -N to nitrogen gas (N2). Of the many factors that are known to affect the SND process, DO concentration is one of the most important. The establishment
100 COD + NH 4-N
80
COD + NH 4-N
o
5±2 C
TN
Removal efficiency (%)
Removal efficiency (%)
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TN
TN
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100
Removal efficiency (%)
Removal efficiency (%)
COD + NH 4-N
80
8
pH
7
8
pH
9
80 60 40 20 0
6
7
8
pH
Fig. 5. Removal of COD, NHþ 4 and TN under different pH values.
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(a) 3.0
(b) 300
o
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2.5
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DO (mg/L)
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o
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1
2
3
4
5
6
7
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0
1
2
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6
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8
Time (h)
Fig. 6. The time courses of DO concentration and ORP under different temperatures.
of both aerobic and anoxic conditions inside a single reactor as a result of DO concentration gradients arising from diffusional limitations inside the activated sludge flocs was essential for the SND process (Liu et al., 2010). The uneven distribution of DO inside the biomass allows simultaneous proliferation of nitrifying and denitrifying bacteria (Chiu et al., 2007). An anoxic zone and an aerobic zone are required in sludge flocs to form the essential environment for SND. Under low DO, diffusional limitations may create an anoxic zone within the biological floc where denitrification can take place. Thus, low DO concentration should be maintained for the accomplishment of SND. Baeza et al. (2002) found that the aerobic reactors using the lowest DO setpoint resulted in the highest removal efficiency; however, the effluent contained an increased proportion of NHþ 4 -N as low DO concentration hinder nitrification but substantially favored denitrification. As DO can penetrate into sludge flocs, high DO concentration reduced the anoxic zone in the inner part of flocs and hence hindered the denitrification. Besides, denitrifiers would switch their electron acceptors from nitrate to oxygen when excessive DO was provided, thereby cease to denitrify. Liu et al. (2010) reported that the denitrification decreased dramatically when the DO concentration was higher than 1.0 mg/ L and inhibition of DO on denitrification was more significant in the incomplete nitrification process. Hence, an appropriate DO level is vital for the SND reactor to reach equilibrium between the nitrification and denitrification processes. Moreover, as shown in Fig. 6, reducing aeration intensity could minimize energy consumption as high DO concentration was insignificant for the removal of organics and the nitrogen compounds under low temperatures. Similar ORP variations were observed under different temperature conditions. The bending point indicated the use up of oxides (NO 2 and NO3 ) and thus absolute ORP can be used as a real-time control parameter for SND process (Wu et al., 2011). After the appearance of the ORP curve bending point, the DO concentration could be reduced by turning down the aeration system, which could avoid the waste of energy for aeration and limit the DO concentration to an appropriate level for better SND efficiency.
3.7. Variations of SVI and EPS amount under different temperatures The average SVI value and EPS amount of 30 d under each temperature were listed in Table 2. The SVI value and the EPS amount decreased with the promotion of temperature. Under 10 ± 2 °C, the SVI increased from 190 on the 17th day to 230 on the 30th day, leading to sludge bulking and poor settleability. While at 5 ± 2 °C,
Table 2 SVI and EPS amount under different temperatures. Temperature (°C) 5±2 10 ± 2 20 ± 2 30 ± 2
SVI (mL/g)
EPS (mg/L)
230.94 210.63 199.19 163.81
235.2 174.5 84.2 34.5
the SVI increased from 250 to 300 and serious sludge bulking and extreme poor settleability were observed. In a SBR system treating pulp and paper mill effluent, the SVI value increased with the enhancement of temperature from 25 to 35 °C and the loose floc led to the poor sludge settleability at higher temperatures (Tsang et al., 2007). Contrary results were obtained in the present study as the SVI value increased with the decrease of temperature from 30 to 5 °C. Krishna and van Loosdrecht (1999) found that microorganisms tended to accumulate more poly-b-hydroxybutyrate (PHB) at lower temperatures. In the present study, higher SVI value was in coincidence with the greater production of EPS when the temperature declined. It was thus presumed that the absorbed carbon substrate or its uptake in a form of PHB was converted to EPS rather than complete mineralization as a result of the low microbial metabolic activity under low temperatures. EPS are one of the representative components of the flocs and are considered to have significant influences on the physicochemical properties of microbial aggregates, including structure, surface charge, flocculation, settling properties, dewatering properties, and adsorption ability. Yang and Li (2009) also found that EPS had a negative effect on the sludge settleability as an increase in EPS content may bring more bound water into the aggregates, and therefore produce highly porous flocs with a low density. Another possible reason for the increase of SVI at lower temperatures was the low density of the sludge flocs due to the rise of gas phase proportion within these flocs resulted from the increased oxygen solubility. Moreover, the settling velocity of the sludge flocs decreased as a result of the increased viscosity of activated sludge at low temperatures, which complicated the separation of treated wastewater and the activated sludge. As a result, the effluent was turbid with small flocs. 3.8. Temperatures oscillations and the counteracts Temperature oscillations in wastewater treatment system were common as a result of seasonal transients. Temperature not only affects the metabolic activities of the microbial population but also
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influences the gas-transfer rates and the settling characteristics of activated sludge. In general, the rate of biochemical reactions and of substrate transfer processes increases with higher temperature. Liu et al. (2010) indicated that when the temperature slipped below 11 °C, the nitrification efficiency decreased sensitively even the DO is enough for the nitrification and the relatively high temperatures (>12 °C) can improves the occurrence of SND. However, the solubility of oxygen decreases in the mixed liquor as temperature increased. The optimum temperatures for common bacterial activity in activated sludge processes were thought to be 25– 35 °C (Krzeminski et al., 2012). When the temperature dropped to about 5 °C, the autotrophic nitrifying bacteria practically ceased functioning and at 2 °C even the chemo-heterotrophic bacteria mineralizing carbonaceous material became essentially dormant (Metcalf and Eddy Inc., 2003). The microbial community and their cell membrane compositions also changed at various temperatures. Thus, temperature generates complicated effects on biochemical reactions. The operational parameters should be adjusted to minimize the influences of temperatures. The optimal operational parameters for each temperature conditions were summarized in Table 3. The SND efficiencies were calculated as described by Ding et al. (2011) and average values were presented, which were comparable as reported by Ding et al. (2011). The prolonged reaction time was beneficial for SND and nitrogen removal in the SBR under low temperatures. Nevertheless, there was a certain threshold of reaction time for each temperature as the further extension of reaction time was insubstantial for performance improvement, only added operational cost for aeration. The suitable influent C/N ratio also varied at different temperatures. However, the discrepancy of SND efficiency at distinct C/N ratio was not remarkable, suggesting that the adjustment of C/N ratio to reduce the impacts of temperature was not an effective option. Longer SRT was required to strengthen nitrification as a result of higher biomass concentrations and the retention of more slow growing microorganisms at low temperatures. Initial pH requirement for different temperatures were similar as neutral to alkali conditions were beneficial for the conversion of organics and the nitrification and denitrification of the nitrogen compounds. An appropriate DO level was demanded to reach equilibrium between the nitrification and denitrification processes in the SBR–SND system. The aeration intensity could be reduced for minimizing energy consumption as the resulted high DO concentration is insignificant for pollutants removal improvements under low temperatures. The ORP, DO concentration, SVI value and EPS amount changed with the temperature alterations, which further led to the variation of the system performance. Previous study suggested that alternative conditions could be utilized to optimize the performance of the existing biological wastewater treatment systems. In an A2/O system, by the dynamic modification of its internal and external recycle flow-rates, the DO setpoint and stirring rates through a distributed control system supervised by a Knowledge Based Expert System (KBES), the amount of nitrogen removed was increased by 11% as compared with the usual operating conditions (Baeza et al., 2002). An intelliTable 3 Optimum operational parameters and nitrogen removal performances under different temperatures. Temperature (°C)
5±2
10 ± 2
20 ± 2
30 ± 2
Reaction time (h) C/N ratio SRT (d) pH TN removal efficiency (%) SND efficiency (%)
8 8 20 8 34 84.6
7 8 15 8 62 88
7 10 10 8 70.9 95.6
6 10 10 8 85.5 97.7
gent controlling system (ICS) controlled sequencing batch biofilm reactor for domestic sewage treatment had shortened the aeration time by 56% and obtained advanced performance (SND efficiency reached 98%) when compared with conventional SBBR (Ding et al., 2011). The optimum fraction of aeration time of a nitrifying SBR was rapidly determined by an algorithm and the reaction phase duration and the operational cost were thus minimized with acceptable effluent quality (Katsogiannis et al., 1999). Real-time modification of process parameters should be performed for better performance and lower operational cost, rather than operating the biological wastewater treatment system under fixed conditions. 4. Conclusions The performance of a SBR–SND was investigated under 5–30 °C. There was a threshold value for aeration time at different temperatures, and extended aeration was beneficial for low temperature conditions but further extension was insubstantial for performance improvement. The adjustment of C/N ratio in a range of 5–12 to offset the impacts of temperature was not a remarkable option. Prolonged SRT favored the performance under low temperatures but was detrimental at high temperatures. The operational and control of these process parameters could alleviate the influences of temperature oscillations, suggesting the significance of real-time modification. Acknowledgements Authors would like to thank the Open Foundation of State Key Laboratory of Urban Water Resource and Environment (No. QA201013), the Science and Technology Development Program of Jilin Province (No. 201101108 and 20110405), the ‘‘Twelfth FiveYear Plan’’ Science and Technology Research Projects of Jilin Province (No. 2012-95) and the Jilin City Planning Project of Science and Technology (201132402) for their financial supports. References Antoniou, P., Hanilton, J., Koopman, B., Jain, R., Holloway, B., Lyberatos, G., Sboronos, S.A., 1990. Effect of temperature and pH on the effective maximum specific growth rate of nitrifying bacteria. Water Res. 24, 97–101. APHA, AWWA, WPCP, 2005. Standard Methods for the Examination of Water and Wastewater. In: Clesceri, L.S., Greenberg, A.E., Eaton, A.D. (Eds.), 21st ed. 21st ed. American Public Health Association, Washington, DC, USA. Baeza, J.A., Gabriel, D., Lafuente, J., 2002. Improving the nitrogen removal efficiency of an A2/O based WWTP by using an on-line knowledge based expert system. Water Res. 36, 2109–2123. Chiemchaisri, C., Yamamoto, K., 1993. Biological nitrogen removal under low temperature in a membrane separation bioreactor. Water Sci. Technol. 28, 325– 333. Chiu, Y.C., Lee, L.L., Chang, C.N., Chao, A.C., 2007. Control of carbon and ammonium ratio for simultaneous nitrification and denitrification in a sequencing batch bioreactor. Int. Biodeterior. Biodegrad. 59, 1–7. Choi, E., Rhu, D., Yun, Z., Lee, E., 1998. Temperature effects on biological nutrient removal system with weak municipal wastewater. Water Sci. Technol. 37, 219– 226. Ding, D., Feng, C., Jin, Y., Hao, C., Zhao, Y., Suemura, T., 2011. Domestic sewage treatment in a sequencing batch biofilm reactor (SBBR) with an intelligent controlling system. Desalination 276, 260–265. Fontenot, Q., Bonvillain, C., Kilgen, M., Boopathy, R., 2007. Effects of temperature, salinity, and carbon: nitrogen ratio on sequencing batch reactor treating shrimp aquaculture wastewater. Bioresour. Technol. 98, 1700–1703. He, S., Xue, G., Wang, B., 2009. Factors affecting simultaneous nitrification and denitrification (SND) and its kinetics model in membrane bioreactor. J. Hazard. Mater. 168, 704–710. Hocaoglu, S.M., Insel, G., Cokgor, E.U., Orhon, D., 2011a. Effect of low dissolved oxygen on simultaneous nitrification and denitrification in a membrane bioreactor treating black water. Bioresour. Technol. 102, 4333–4340. Hocaoglu, S.M., Insel, G., Cokgor, E.U., Orhon, D., 2011b. Effect of sludge age on simultaneous nitrification and denitrification in membrane bioreactor. Bioresour. Technol. 102, 6665–6672. Holakoo, L., Nakhla, G., Bassi, A.S., Yanful, E.K., 2007. Long term performance of MBR for biological nitrogen removal from synthetic municipal wastewater. Chemosphere 66, 849–857.
J. Guo et al. / Bioresource Technology 133 (2013) 59–67 Holman, J.B., Wareham, D.G., 2005. COD, ammonia and dissolved oxygen time profiles in the simultaneous nitrification/denitrification process. Biochem. Eng. J. 22, 125–133. Ilies, P., Mavinic, D.S., 2001. The effect of decreased ambient temperature on the biological nitrification and denitrification of a high ammonia landfill leachate. Water Res. 35, 2065–2072. Johnson, K., van Geest, J., Kleerebezem, R., van Loosdrecht, M.C.M., 2010. Short- and long-term temperature effects on aerobic polyhydroxybutyrate producing mixed cultures. Water Res. 44, 1689–1700. Karkman, A., Mattila, K., Tamminen, M., Virta, M., 2011. Cold temperature decreases bacterial species richness in nitrogen-removing bioreactors treating inorganic mine waters. Biotechnol. Bioeng. 108, 2876–2883. Katsogiannis, A.N., Kornaros, M.E., Lyberatos, G.K., 1999. Adaptive optimization of a nitrifying sequencing batch reactor. Water Res. 33, 3569–3576. Kayranli, B., Ugurlu, A., 2011. Effects of temperature and biomass concentration on the performance of anaerobic sequencing batch reactor treating low strength wastewater. Desalination 278, 77–83. Komorowska-Kaufman, M., Majcherek, H., Klaczyn´ski, E., 2006. Factors affecting the biological nitrogen removal from wastewater. Process Biochem. 41, 1015–1021. Krishna, C., van Loosdrecht, M.C.M., 1999. Effect of temperature on storage polymers and settleability of activated sludge. Water Res. 33, 2374–2382. Krzeminski, P., Iglesias-Obelleiro, A., Madebo, G., Garrido, J.M., vander Graaf, J.H.J.M., van Lier, J.B., 2012. Impact of temperature on raw wastewater composition and activated sludge filterability in full-scale MBR systems for municipal sewage treatment. J. Membr. Sci. 423–424, 348–361. Lan, C.J., Kumar, M., Wang, C.C., Lin, J.G., 2011. Development of simultaneous partial nitrification, anammox and denitrification (SNAD) process in a sequential batch reactor. Bioresour. Technol. 102, 5513–5519. Liang, Z., Li, W., Yang, S., Du, P., 2010. Extraction and structural characteristics of extracellular polymeric substances (EPS), pellets in autotrophic nitrifying biofilm and activated sludge. Chemosphere 81, 626–632. Liu, Y., Shi, H., Xia, L., Shi, H., Shen, T., Wang, Z., Wang, G., Wang, Y., 2010. Study of operational conditions of simultaneous nitrification and denitrification in a
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Carrousel oxidation ditch for domestic wastewater treatment. Bioresour. Technol. 101, 901–906. Metcalf & Eddy Inc. 2003. Wastewater Engineering: Treatment and Reuse, fourth ed. McGraw-Hill, New York. Rong, Q., Yang, K., Yu, Z., 2007. Treatment of coke plant wastewater by SND fixed biofilm hybrid system. J. Environ. Sci. 19, 153–159. Seifi, M., Fazaelipoor, M.H., 2012. Modeling simultaneous nitrification and denitrification (SND) in a fluidized bed biofilm reactor. Appl. Math. Model. 36, 5603–5613. Sudarno, U., Winter, J., Gallert, C., 2011. Effect of varying salinity, temperature, ammonia and nitrous acid concentrations on nitrification of saline wastewater in fixed-bed reactors. Bioresour. Technol. 101, 5665–5673. Tsang, Y.F., Hua, F.L., Chua, H., Sin, S.N., Wang, Y.J., 2007. Optimization of biological treatment of paper mill effluent in a sequencing batch reactor. Biochem. Eng. J. 34, 193–199. Wan, J., Bessière, Y., Spérandio, M., 2009. Alternating anoxic feast/aerobic famine condition for improving granular sludge formation in sequencing batch airlift reactor at reduced aeration rate. Water Res. 43, 5097–5108. Willers, H.C., ten Have, P.J.W., Derikx, P.J.L., Arts, M.W., 1993. Temperaturedependency of nitrification and required anoxic volume for denitrification in the biological treatment of veal calf manure. Bioresour. Technol. 43, 47–52. Wu, C., Peng, Y., Wang, S., Li, X., Wang, R., 2011. Effect of sludge retention time on nitrite accumulation in real-time control biological nitrogen removal sequencing batch reactor. Chin. J. Chem. Eng. 19, 512–517. Xia, S., Li, J., Wang, R., 2008. Nitrogen removal performance and microbial community structure dynamics response to carbon nitrogen ratio in a compact suspended carrier biofilm reactor. Ecol. Eng. 32, 256–262. Yang, S., Li, X., 2009. Influences of extracellular polymeric substances (EPS) on the characteristics of activated sludge under non-steady-state conditions. Process Biochem. 44, 91–96. Yang, S., Yang, F., 2011. Nitrogen removal via short-cut simultaneous nitrification and denitrification in an intermittently aerated moving bed membrane bioreactor. J. Hazard. Mater. 195, 318–323.
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Bioresource Technology journal homepage: www.elsevier.com/locate/biortech
The regulation and control strategies of a sequencing batch reactor for simultaneous nitrification and denitrification at different temperatures Jingbo Guo a,⇑, Lanhe Zhang b, Wei Chen a, Fang Ma c, Honglei Liu b, Yu Tian b a
School of Civil and Architecture Engineering, Northeast Dianli University, Jilin 132012, China School of Chemical Engineering, Northeast Dianli University, Jilin 132012, China c State Key Lab of Urban Water Resource and Environment, School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China b
h i g h l i g h t s " Prolonged aeration favored the SBR–SND system under low temperatures. " Further extension of aeration was insubstantial, only increased operational cost. " C/N ratio adjustment to offset the temperature impacts was ineffective. " Long SRT lessened low temperature effect but reduced N removal at high temperature. " ORP, DO, SVI and EPS changed with temperature and led to performance variation.
a r t i c l e
i n f o
Article history: Received 23 November 2012 Received in revised form 5 January 2013 Accepted 5 January 2013 Available online 22 January 2013 Keywords: Simultaneous nitrification and denitrification Sequencing batch reactor Temperature Regulation and control strategies
a b s t r a c t The performance of a sequencing batch reactor for simultaneous nitrification and denitrification (SBR– SND) was investigated under 5–30 °C and strategies against temperature influences were proposed. Aeration of 8, 7, 7 and 6 h were sufficient for 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C, respectively. Further extension was insubstantial, only increased the aeration cost. The adjustment of C/N ratio to offset the temperature impacts was not remarkable. Prolonged sludge retention time lessened the influences of low temperature but deteriorated the system at high temperature. The oxidation reduction potential, the dissolved oxygen concentration, the sludge volume index and the extracellular polymeric substances amount changed with temperature alterations and thus affected the system performance. In conclusion, measures should be taken for temperature oscillations and the regulation and control of the operational parameters could alleviate the influences of temperature on the performances of the SBR–SND system. Ó 2013 Elsevier Ltd. All rights reserved.
1. Introduction The adverse environmental impacts of nitrogenous compounds include promotion of eutrophication, toxicity to aquatic organisms and depletion of dissolved oxygen (DO) in receiving water bodies due to bacterial oxidation of ammonia to nitrate (He et al., 2009). Nitrogen removal to an appropriate concentration is important and compulsory before wastewaters are being discharged into the environment (Seifi and Fazaelipoor, 2012). Biological treatment is one of the most economical processes for nitrogen removal. Biological nitrogen removal (BNR) from wastewaters is usually accomplished through sequential nitrification and denitrification processes, i.e., conventional nitrification–denitrification process, by which nitrification and denitrification are performed in separate compartments (Lan et al., 2011). In the nitrification compartment, ⇑ Corresponding author. Tel./fax: +86 432 62834526. E-mail address: [email protected] (J. Guo). 0960-8524/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.biortech.2013.01.026
aerobic autotrophic nitrifiers oxidize NHþ 4 —N to NO2 —N and then NO3 -N; in the denitrification compartment, NO3 —N is reduced to NO 2 —N and then to gaseous nitrogen compounds by heterotrophic denitrifying bacteria under anoxic conditions (Wan et al., 2009). Studies showed that nitrification and denitrification could occur concurrently in a single reactor under aerobic conditions, which is often referred to as the simultaneous nitrification and denitrification (SND) process (Yang and Yang, 2011). The SND process represents a significant advantage over the conventional separated nitrification and denitrification processes. First, SND system eliminates the serial operation of two separate tanks and thus requires a smaller footprint and simpler operating procedures (Yang and Yang, 2011). Favorable nitrogen removal efficiency could be achieved at lower C/N ratios as more carbon substrates are available for denitrifers in the SND system (Hocaoglu et al., 2011a). It is estimated that the SND process utilizes 22–40% less carbon source and reduces sludge yield by 30% when compared with the conventional BNR system (Seifi and Fazaelipoor, 2012). The SND
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could be accomplished under neutral pH level with less demand for alkalinity because alkalinity is consumed during nitrification but produced during denitrification (Rong et al., 2007). Also, oxygen consumption may be greatly reduced because NO 3 -N can be utilized to the fullest extent as the alternate electron acceptor and thus low DO concentration (0.10–1.00 mg/L) is sufficient for SND reaction (Chiu et al., 2007; Hocaoglu et al., 2011a,b). Therefore, to take full advantages of this process, factors including carbon and ammonium ratio (C/N) (Chiu et al., 2007), sludge retention time (SRT) (Wu et al., 2011), DO concentration (Hocaoglu et al., 2011a) and temperature (Ilies and Mavinic, 2001) as well as other factors (He et al., 2009) on the performances of the SND process had been widely investigated. As an important environmental factor of the biological wastewater treatment process, temperature not only influences the metabolic activities of the microbial population, but also has a profound effect on such factors as gas-transfer rates and the settling characteristics of the biological solids (Metcalf and Eddy Inc., 2003). Biological reaction rates increase with rising temperature until an optimal temperature is reached; above the optimal temperature, enzymatic proteins denature and the rates decrease. The nitrification and denitrification of a high ammonia landfill leachate were greatly inhibited at 10 °C (Ilies and Mavinic, 2001). Temperature and NHþ 4 -N removal rate were positive correlated with temperature in an interval of 5–20 °C (Willers et al., 1993). The nitrification rate was an increasing function of temperature in the range of 15–25 °C (Antoniou et al., 1990). Maximum nitrogen removal of a shrimp aquaculture wastewater with a SBR was obtained in the presence of 2.8–4% salinity and in a temperature range of 22–37 °C (Fontenot et al., 2007). The temperatures of municipal wastewater vary in the range of 10–30 °C and may fall below 10 °C in cold areas, for example, the North China. To date, there are limited reports about the influences of temperature on the performance of the SND system and the corresponding coping measures. In the present study, the performance of a sequencing batch reactor based simultaneous nitrification and denitrification (SBR–SND) system was investigated under different temperatures feeding with synthetic wastewater. The changes of oxidation reduction potential (ORP), DO concentration, the sludge volume index (SVI) value and the amount of extracellular polymeric substances (EPS) under different temperatures were detected. The regulation and control strategies were proposed to achieve relatively high nitrogen removal efficiency, low costs and easy maintenance. The objective of this study is to investigate the influences of temperature on the performance of the SBR–SND system and provide useful guidance for its operation in accordance with temperature oscillations, rather than following the stereotype without any alternation.
2. Methods 2.1. The lab-scale SBR and the operating conditions A lab-scale SBR with dimensions of 30 10 cm (H D) and a working volume of 2.5 L was fabricated with polymethyl methacrylate. Samples were collected from the three sampling outlets located on the SBR side. Sludge discharging outlets was located at the bottom of the SBR. A complete mixing inside the SBR was ensured by suspending the mixed liquor with a mechanical stirrer at a rate of 100 rpm. Aeration was provided by pressurized air passing through long stone diffusers and the initial DO concentration was controlled at 0.5–1.0 mg/L by air flow valve. Incubator was used to provide different temperatures. Sludge collected from the aerobic pond of a municipal wastewater treatment plant (Jilin, China) was used as inoculum. The sludge was elutriated with clean water
three times to remove fine suspended solids that may interfere with the analyses. Then the sludge was added to SBR at a final concentration of approximately 1500 mg/L in the synthetic wastewater. One typical cycle of the SBR contained four procedures: feed (instantaneous), reaction (7.5 h), settling (0.5 h), decant (instantaneous, 80% water exchange, 2 L), idle (4 h). 2.2. Synthetic wastewater Synthetic wastewater was used for investigating the SND performance at different temperatures. The composition was as follows (per liter): CH3COONa, 0.6–1.5 g; KH2PO3, 20 mg; MgSO4, 60 mg; CaCl2, 360 mg; FeSO4, 3 mg; NH4Cl, 35–50 mg; trace element solution, 0.5 mL. The trace element solution contained the following chemicals (per liter): FeCl36H2O, 1.5 g; H3BO3, 0.15 g; CuSO45H2O, 0.03 g; KI, 1.18 g; MnCl24H2O, 0.12 g; ZnSO47H2O, 0.12 g; CoCl2, 0.15 g. The characteristics of the synthetic wastewater are summarized in Table 1. 2.3. Experimental plan During the start-up period, the SBR was operated with a hydraulic retention time (HRT) of 10 h and a sludge retention time (SRT) of 20 d at 25 ± 2 °C. The influent COD concentration was approximately 450 mg/L and the TN concentration was gradually increased from 10 to 50 mg/L. After the steady-state was achieved as demonstrated by efficient and stable COD and TN removal efficiencies, the SBR was operated for an additional 20 d at 25 ± 2 °C. Then the temperature of the incubator where the SBR placed was set at 5 ± 2 °C. The effluent quality was monitored until the steady-state was achieved. Then an average value of effluent quality from seven serial cycles was presented to reveal the necessary aeration time at 5 ± 2 °C. The performances of the SBR–SND system under different C/N ratios at the same temperature were then investigated applying the necessary aeration time. Then the SRT was controlled through direct removal of sludge from the bioreactor (1/5, 1/10, 1/15 and 1/20 of the bioreactor volume) on daily basis, to maintain SRT of 5, 10, 15 and 20 d respectively for exploring the influences of SRT on the SBR–SND system. After that, the influences of the pH value were investigated by adjusting the influent pH in a range of 4.8–9.7. The performances of the SBR–SND system and the corresponding SVI and EPS amount was detected during 30 d under 5 ± 2 °C when the SND–SBR system were operated under the optimal parameters. Meanwhile, the variation of DO and ORP value in a typical cycle were detected. Then the temperature of the incubator was successively set at 10 ± 2, 20 ± 2, 30 ± 2 °C and the same experimental procedures were conducted under each temperature. Two turnovers of the SRT (20 d) were operated to reach the steady-state for each temperature change. Finally, data obtained from different temperatures were collected and compared to reveal the characteristics of the SBR–SND system under various temperatures and the corresponding regulation and control strategies. 2.4. Analytic methods The concentration of COD, NHþ 4 —N, NO2 —N, NO3 —N, TN and mixed liquor suspended solids (MLSS) were measured according
Table 1 The characteristics of the synthetic wastewater. Parameters
COD
TN
NHþ 4 -N
NO 3 -N
NO 2 -N
Range (mg/L)
100–500
40–42
40–42
0–0.5
0–0.5
J. Guo et al. / Bioresource Technology 133 (2013) 59–67
to Standard Methods for the Examination of Water and Wastewater (APHA, AWWA, WPCP, 2005). Temperature, pH, ORP and DO were analyzed online with a WTW pH/ORP sensor (WTW, pH 3310) and a DO sensor (WTW, OXi340i). The extraction and quantification of EPS were performed according to the protocol described by Liang et al. (2010). 3. Results and discussions 3.1. Start-up of the SBR The reactors became stable after being acclimated for 30 d. As shown in Fig. 1, the COD and TN removal efficiencies were up to 90% and 85%, respectively. The effluent TN concentration was 3– 8.56 mg/L when the influent TN concentration varied from 10 mg/L to 53.5 mg/L. SND was accomplished as the nitrification and denitrification occurred simultaneously in a single reactor under identical conditions. 3.2. Adjustment of reaction time under different temperatures The aeration duration was set as 10 h and the effluent COD, TN and NHþ 4 -N concentration was detected at hourly intervals when the influent NHþ 4 -N concentration was 40–42 mg/L. The data presented in Fig. 2 was an average value of seven serial cycles at the steady state of each temperature. The removal of COD could be divided into three phase. At 5 ± 2 and 10 ± 2 °C, more than 50% COD was removed within the initial 4 h; the COD removal efficiency reduced dramatically during the following 2 h; and then the effluent COD concentration hardly changed thereafter. At 20 ± 2 and
100
600 500
80
400 60
Inf. COD Eff. COD COD removal efficiency
300
40
200 20
100 0
0
5
10
15
20
25
30
COD removal efficiency (%)
COD concentration (mg/L)
(a)
0
Time (d) 60
100
50
80
40
Inf. TN Eff. TN TN removal efficiency
30
60
40 20 20
10 0
0
5
10
15
20
25
30
0
Time (d) Fig. 1. The removal of COD (a) and TN (b) during start-up period.
TN removal efficiency (%)
TN concentration (mg/L)
(b)
61
30 ± 2 °C, up to 80% COD was removed within the initial 3 h; the removal rate decreased gradually in the following 2 h; little changes of COD concentration were observed thereafter. The effluent NHþ 4 -N and TN stabilized after 8, 7, 7 and 6 h operation when the respective temperatures were 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. Thus, the nitrogen removal efficiency increased with the elevation of temperature and the prolonged reaction time was beneficial for the SBR operated under low temperatures. Microorganisms could store carbon substrates and then served as carbon sources for denitrification and heterotrophic nitrification (Ding et al., 2011). This could explain the higher decreasing rate of COD at the initial stage than the nitrogen compounds and then the COD concentration stabilized, while the TN and NHþ 4 -N concentrations deceased at higher rates. There was a discrepancy between the effluent NHþ 4 -N and TN concentration, suggesting the accumulation of NO 3 -N and/or NO 2 -N. At 5 ± 2 °C, this discrepancy was small as the nitrification and the denitrification were greatly inhibited by low temperature and the majority of the added nitrogen compound was discharged without decomposition. This discrepancy decreased with the elevation of temperature from 10 ± 2 to 20 ± 2 °C and then to 30 ± 2 °C as the equilibrium of the nitrification and denitrification rates was improved. The effect of temperature on NO 3 -N and/or NO -N accumulation was probably relative to the growth charac2 teristics of nitrifiers and denitrifiers, namely, denitrifiers were more sensitive than nitrifiers to low temperature. This was in accordance with the previous study that the denitrification was a key step to regulate the overall nutrient removal efficiencies at low temperatures (Choi et al., 1998). S According to Michaelis–Menten equation V ¼ KVmmþS and ArrheEa nius equation V m ¼ AexpRT , high temperature will improve the biodegradation rate as a result of the enhanced microbial enzyme activity (Johnson et al., 2010). An increment of 10 °C from 22.5 to 32.5 °C resulted in an improvement of the ammonium oxidation rate (AOR) and nitrite oxidation rate (NOR) by a factor of 1.36 and 1.26, respectively; the AOR and NOR decreased to 44% and 55%, of the initial rates at decreasing temperatures from 22.5 to 12.5 °C (Sudarno et al., 2011). As a counteraction of the low temperature wastewater treatment, the biodegradation duration was frequently extended by prolonging hydraulic retention time (HRT) or aeration time (Kayranli and Ugurlu, 2011). Chiemchaisri and Yamamoto (1993) found that by increasing aeration time in operational cycle of a membrane separation bioreactor for on-site domestic wastewater treatment, the nitrification could be recovered at low temperatures as a result of the enhanced oxygen supply. Low temperature not only influences the metabolic activities of nitrifiers and denitrfiers, but also affects the bacterial community and the species richness. Karkman et al. (2011) identified the bacteria in a bioreactor treating inorganic mine waters under 5 and 10 °C and found that the dominant nitrifiers were distinct at different temperature; meanwhile, as the temperature increased from 5 to 10 °C, the operational taxonomic units (OTUs) elevated from lower than 150 to higher than 300, demonstrating that the bacterial species richness enhanced with the elevation of temperature. In the present case, 8, 7, 7 and 6 h were sufficient when the respective temperatures were 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. The further extension of reaction time was insubstantial for performance improvement, especially at low temperature conditions, only increased the operational cost for aeration. Besides, shorter reaction time could provide more degradable organic substrates as reducing power for denitrification in a SND system; otherwise, the readily-biodegradable COD was oxidized by oxygen in the bioreactor (Liu et al., 2010). Thus, for optimizing SND efficiency and reducing operational cost, an appropriate reaction time should be determined under different temperature conditions and the pro-
J. Guo et al. / Bioresource Technology 133 (2013) 59–67
50
300 30 200
20
100
+
10
COD concentration (mg/L)
NH4 -N and TN concentration (mg/L)
Eff. TN Eff. COD
40
400
+
5±2 oC 0
NH4 -N and TN concentration (mg/L)
+
Eff. NH4 -N
0
2
4
6
8
10
0
+
Eff. NH4 -N Eff. TN Eff. COD
40
300 30 200
20
10
0
100
10±2 oC
0
2
Time (d)
200
20
100
10
20±2 oC 2
4
6
8
10
0
Time (d)
+
300 30
COD concentration (mg/L)
400
NH4 -N and TN concentration (mg/L)
+
NH4 -N and TN concentration (mg/L)
Eff. TN Eff. COD
0
6
8
50
+
Eff. NH4 -N
0
4
10
0
Time (d)
50
40
400
COD concentration (mg/L)
50
+
Eff. NH4 -N Eff. TN Eff. COD
40
400
300 30 200
20
100
10
COD concentration (mg/L)
62
30±2 oC 0
0
2
4
6
8
10
0
Time (d)
Fig. 2. The time courses of effluent NHþ 4 , TN and COD concentrations under different temperatures.
longed reaction time was not a radical solution for low temperature wastewater treatment. 3.3. Adjustment of influent C/N ratio under different temperatures In SND system, less carbon was demanded as nitrification and denitrification was accomplished in a single reactor when compared with the conventional BNR processes. As the autotrophic nitrification is generally slower than the heterotrophic metabolism, SND requires a degradable organic substrate to provide reducing power for heterotrophic nitrification and denitrification processes (Liu et al., 2010). Nevertheless, competition between autotrophic and heterotrophic microorganisms would be induced at high organic loading rate (OLR). In a sequencing batch airlift reactor, the competition for oxygen between heterotrophic and autotrophic bacteria became critical at OLR higher than 2 kg COD/(m3 d) (Wan et al., 2009). Hence, the C/N ratio should be adjusted to achieve the equilibrium between nitrification and denitrification processes. As shown in Fig. 3, there was an optimal C/N ratio for the removal of organics and the nitrogen compounds at different temperatures. The highest COD, NHþ 4 -N and TN removal efficiencies at 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C were obtained when the influent C/N ratios were 8, 8, 10, 10, respectively, indicating that C/N ratio requirements varied at different temperatures. Low C/N ratio resulted in a rapid carbon deficit, causing an unbalanced SND process in SBR. Complete removal of NHþ 4 -N was achieved when the initial C/N was adjusted to 11.1 at 25 ± 2 °C (Chiu et al., 2007). On the other side, in a compact suspended carrier biofilm reactor (SCBR) developed for SND, the diversity of microbial community structure and the population of nitrifiers was in inverse propor-
tions to C/N ratio (Xia et al., 2008). Therefore, it’s necessary to choose an appropriate C/N ratio according to the types of SND process being adopted and wastewater being treated. However, the nitrogen removal efficiencies obtained at various influent C/N ratios of a certain temperature was not remarkable, suggesting that the adjustment of C/N ratio to offset the impacts of temperature was not an effective option. 3.4. Adjustment of sludge retention time (SRT) under different temperatures As shown in Fig. 4, the best performances were obtained with a 20, 15, 10 and 10 d SRT at 5 ± 2, 10 ± 2, 20 ± 2 and 30 ± 2 °C. The removal efficiency of NHþ 4 -N was higher than that of TN at all condi tions, indicating the accumulation of NO 3 -N and/or NO2 -N. The biomass concentration in the biological reactor is a function of the SRT and the different biomass concentrations resulted from the variations of SRT was crucial for the mass transfer conditions between DO, nitrogen compounds and biomass fractions responsible for nitrifications and denitrification (Hocaoglu et al., 2011b). Long SRT could reinforce nitrification as a result of biomass enhancement and better retention of slow growing microorganisms. In a membrane bioreactor for the treatment of synthetic municipal wastewater, the highest total nitrogen removal and nitrogen removal via SND were achieved with a SRT of 40 d under 16 ± 1 °C and a DO concentration of 0.5–1.5 mg/L (Holakoo et al., 2007) and decreased with a SRT of 20 d. Holakoo et al. (2007) indicated the deteriorated SND with decreased SRT primarily attributed to a reduction of the floc size, which limited denitrification as the overall anoxic volume fraction of the flocs reduced. SRT extension was one of the common measures to sustain the biolog-
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100
100 80
TN
60 40 20 0
C/N=10
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(%)
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20 0
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60
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Time (d)
Fig. 3. Removal of COD, NHþ 4 and TN under different effluent C/N ratios.
100
100 COD + NH4 -N
5±2 C
80
TN
Removal efficiency (%)
Removal efficiency (%)
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o
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SRT (h) 100
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TN
Removal efficiency (%)
Removal efficiency (%)
COD + NH4 -N
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TN
80
10±2 oC
10
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SRT (h)
20
80
60
40
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SRT (h)
Fig. 4. Removal of COD, NHþ 4 and TN under different SRTs.
20
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ical wastewater treatment performance and stability under low temperatures as increased sludge age could lessen the unfavorable effect of low temperatures (under 15 °C) and stabilize the nitrification process (Komorowska-Kaufman et al., 2006). However, as shown in Fig. 4, the carbon and nitrogen removal efficiencies were not always proportional to the SRT as inflection points were detected at SRT of 15 or 10 d when temperatures were higher than 5 ± 2 °C. Long SRT would accelerate the floc breakup, especially under high temperature conditions (Holakoo et al., 2007). Thus, the extension of SRT in a suitable range could lessen the influences of low temperature but the biological nitrogen capability of the SBR–SND system would be detrimentally affected by longer SRT at higher temperatures. 3.5. Adjustment of pH under different temperatures In the conventional nitrification and denitrification system, the generation of acid compounds during the nitrification process will cause pH reduction; on the contrary, the pH increases in the denitrification process as a result of the alkali formation. The pH was often taken as an indicator of the progressing degree of nitrification and denitrification. In the SND system, the alkalinity generated by the denitrification could partially compensate for that consumed by the nitrification process. As shown in Fig. 5, the best performance of the SBR was obtained with pH 8 at all temperatures, suggesting that neutral to alkali conditions were beneficial for the conversion of organics and the nitrification and denitrification of the nitrogen compounds. This was consistent with the results obtained by He et al. (2009) that the optimum TN removal was detected at around pH 7.2 with an investigated range from 4.8 to 9.7.
3.6. Variation of DO concentration and ORP under different temperatures The saturated DO concentration in water increases with the reduction of temperature and thus the DO concentration in the SBR will vary with temperature fluctuations. The relationship between the saturated DO concentration and the DO concentration detected in the reactor could be represented by the equation of oxygen transfer rate: ddDOt ¼ K La ðDOs DOÞ. The KLa value increases with the elevation of temperature. On the contrary, the saturated DO concentration decreases as the temperature enhances. Therefore, the temperature has reverse effects on the oxygen transfer rate. However, these effects could not be offset and generally the low temperature is beneficial for oxygen transfer. As shown in Fig. 6, the DO concentration showed a slight decrease at the initial stage, while longer durations of the decreasing trend were observed at higher temperatures. Holman and Wareham (2005) demonstrated that the depletion of COD and NHþ 4 -N caused a sudden DO increase, which was also observed in the present study under 20 ± 2 and 30 ± 2 °C. The increase of DO concentration at 5 ± 2 and 15 ± 2 °C emerged earlier than at higher temperatures, which were the joint results of higher saturated DO concentration and lower metabolic activities of microorganisms under low temperatures. A decrease in microbial activity would cause a decrease in oxygen and substrate utilization and resulted in high DO and effluent COD and nitrogen concentration at low temperature conditions. The increase in DO would, in turn, hamper denitrification, which requires anaerobic environments in order to convert NO 3 -N to nitrogen gas (N2). Of the many factors that are known to affect the SND process, DO concentration is one of the most important. The establishment
100 COD + NH 4-N
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Fig. 5. Removal of COD, NHþ 4 and TN under different pH values.
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(a) 3.0
(b) 300
o
5± 2 C o 10 ± 2 C o 20 ± 2 C o 30 ± 2 C
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Fig. 6. The time courses of DO concentration and ORP under different temperatures.
of both aerobic and anoxic conditions inside a single reactor as a result of DO concentration gradients arising from diffusional limitations inside the activated sludge flocs was essential for the SND process (Liu et al., 2010). The uneven distribution of DO inside the biomass allows simultaneous proliferation of nitrifying and denitrifying bacteria (Chiu et al., 2007). An anoxic zone and an aerobic zone are required in sludge flocs to form the essential environment for SND. Under low DO, diffusional limitations may create an anoxic zone within the biological floc where denitrification can take place. Thus, low DO concentration should be maintained for the accomplishment of SND. Baeza et al. (2002) found that the aerobic reactors using the lowest DO setpoint resulted in the highest removal efficiency; however, the effluent contained an increased proportion of NHþ 4 -N as low DO concentration hinder nitrification but substantially favored denitrification. As DO can penetrate into sludge flocs, high DO concentration reduced the anoxic zone in the inner part of flocs and hence hindered the denitrification. Besides, denitrifiers would switch their electron acceptors from nitrate to oxygen when excessive DO was provided, thereby cease to denitrify. Liu et al. (2010) reported that the denitrification decreased dramatically when the DO concentration was higher than 1.0 mg/ L and inhibition of DO on denitrification was more significant in the incomplete nitrification process. Hence, an appropriate DO level is vital for the SND reactor to reach equilibrium between the nitrification and denitrification processes. Moreover, as shown in Fig. 6, reducing aeration intensity could minimize energy consumption as high DO concentration was insignificant for the removal of organics and the nitrogen compounds under low temperatures. Similar ORP variations were observed under different temperature conditions. The bending point indicated the use up of oxides (NO 2 and NO3 ) and thus absolute ORP can be used as a real-time control parameter for SND process (Wu et al., 2011). After the appearance of the ORP curve bending point, the DO concentration could be reduced by turning down the aeration system, which could avoid the waste of energy for aeration and limit the DO concentration to an appropriate level for better SND efficiency.
3.7. Variations of SVI and EPS amount under different temperatures The average SVI value and EPS amount of 30 d under each temperature were listed in Table 2. The SVI value and the EPS amount decreased with the promotion of temperature. Under 10 ± 2 °C, the SVI increased from 190 on the 17th day to 230 on the 30th day, leading to sludge bulking and poor settleability. While at 5 ± 2 °C,
Table 2 SVI and EPS amount under different temperatures. Temperature (°C) 5±2 10 ± 2 20 ± 2 30 ± 2
SVI (mL/g)
EPS (mg/L)
230.94 210.63 199.19 163.81
235.2 174.5 84.2 34.5
the SVI increased from 250 to 300 and serious sludge bulking and extreme poor settleability were observed. In a SBR system treating pulp and paper mill effluent, the SVI value increased with the enhancement of temperature from 25 to 35 °C and the loose floc led to the poor sludge settleability at higher temperatures (Tsang et al., 2007). Contrary results were obtained in the present study as the SVI value increased with the decrease of temperature from 30 to 5 °C. Krishna and van Loosdrecht (1999) found that microorganisms tended to accumulate more poly-b-hydroxybutyrate (PHB) at lower temperatures. In the present study, higher SVI value was in coincidence with the greater production of EPS when the temperature declined. It was thus presumed that the absorbed carbon substrate or its uptake in a form of PHB was converted to EPS rather than complete mineralization as a result of the low microbial metabolic activity under low temperatures. EPS are one of the representative components of the flocs and are considered to have significant influences on the physicochemical properties of microbial aggregates, including structure, surface charge, flocculation, settling properties, dewatering properties, and adsorption ability. Yang and Li (2009) also found that EPS had a negative effect on the sludge settleability as an increase in EPS content may bring more bound water into the aggregates, and therefore produce highly porous flocs with a low density. Another possible reason for the increase of SVI at lower temperatures was the low density of the sludge flocs due to the rise of gas phase proportion within these flocs resulted from the increased oxygen solubility. Moreover, the settling velocity of the sludge flocs decreased as a result of the increased viscosity of activated sludge at low temperatures, which complicated the separation of treated wastewater and the activated sludge. As a result, the effluent was turbid with small flocs. 3.8. Temperatures oscillations and the counteracts Temperature oscillations in wastewater treatment system were common as a result of seasonal transients. Temperature not only affects the metabolic activities of the microbial population but also
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influences the gas-transfer rates and the settling characteristics of activated sludge. In general, the rate of biochemical reactions and of substrate transfer processes increases with higher temperature. Liu et al. (2010) indicated that when the temperature slipped below 11 °C, the nitrification efficiency decreased sensitively even the DO is enough for the nitrification and the relatively high temperatures (>12 °C) can improves the occurrence of SND. However, the solubility of oxygen decreases in the mixed liquor as temperature increased. The optimum temperatures for common bacterial activity in activated sludge processes were thought to be 25– 35 °C (Krzeminski et al., 2012). When the temperature dropped to about 5 °C, the autotrophic nitrifying bacteria practically ceased functioning and at 2 °C even the chemo-heterotrophic bacteria mineralizing carbonaceous material became essentially dormant (Metcalf and Eddy Inc., 2003). The microbial community and their cell membrane compositions also changed at various temperatures. Thus, temperature generates complicated effects on biochemical reactions. The operational parameters should be adjusted to minimize the influences of temperatures. The optimal operational parameters for each temperature conditions were summarized in Table 3. The SND efficiencies were calculated as described by Ding et al. (2011) and average values were presented, which were comparable as reported by Ding et al. (2011). The prolonged reaction time was beneficial for SND and nitrogen removal in the SBR under low temperatures. Nevertheless, there was a certain threshold of reaction time for each temperature as the further extension of reaction time was insubstantial for performance improvement, only added operational cost for aeration. The suitable influent C/N ratio also varied at different temperatures. However, the discrepancy of SND efficiency at distinct C/N ratio was not remarkable, suggesting that the adjustment of C/N ratio to reduce the impacts of temperature was not an effective option. Longer SRT was required to strengthen nitrification as a result of higher biomass concentrations and the retention of more slow growing microorganisms at low temperatures. Initial pH requirement for different temperatures were similar as neutral to alkali conditions were beneficial for the conversion of organics and the nitrification and denitrification of the nitrogen compounds. An appropriate DO level was demanded to reach equilibrium between the nitrification and denitrification processes in the SBR–SND system. The aeration intensity could be reduced for minimizing energy consumption as the resulted high DO concentration is insignificant for pollutants removal improvements under low temperatures. The ORP, DO concentration, SVI value and EPS amount changed with the temperature alterations, which further led to the variation of the system performance. Previous study suggested that alternative conditions could be utilized to optimize the performance of the existing biological wastewater treatment systems. In an A2/O system, by the dynamic modification of its internal and external recycle flow-rates, the DO setpoint and stirring rates through a distributed control system supervised by a Knowledge Based Expert System (KBES), the amount of nitrogen removed was increased by 11% as compared with the usual operating conditions (Baeza et al., 2002). An intelliTable 3 Optimum operational parameters and nitrogen removal performances under different temperatures. Temperature (°C)
5±2
10 ± 2
20 ± 2
30 ± 2
Reaction time (h) C/N ratio SRT (d) pH TN removal efficiency (%) SND efficiency (%)
8 8 20 8 34 84.6
7 8 15 8 62 88
7 10 10 8 70.9 95.6
6 10 10 8 85.5 97.7
gent controlling system (ICS) controlled sequencing batch biofilm reactor for domestic sewage treatment had shortened the aeration time by 56% and obtained advanced performance (SND efficiency reached 98%) when compared with conventional SBBR (Ding et al., 2011). The optimum fraction of aeration time of a nitrifying SBR was rapidly determined by an algorithm and the reaction phase duration and the operational cost were thus minimized with acceptable effluent quality (Katsogiannis et al., 1999). Real-time modification of process parameters should be performed for better performance and lower operational cost, rather than operating the biological wastewater treatment system under fixed conditions. 4. Conclusions The performance of a SBR–SND was investigated under 5–30 °C. There was a threshold value for aeration time at different temperatures, and extended aeration was beneficial for low temperature conditions but further extension was insubstantial for performance improvement. The adjustment of C/N ratio in a range of 5–12 to offset the impacts of temperature was not a remarkable option. Prolonged SRT favored the performance under low temperatures but was detrimental at high temperatures. The operational and control of these process parameters could alleviate the influences of temperature oscillations, suggesting the significance of real-time modification. Acknowledgements Authors would like to thank the Open Foundation of State Key Laboratory of Urban Water Resource and Environment (No. QA201013), the Science and Technology Development Program of Jilin Province (No. 201101108 and 20110405), the ‘‘Twelfth FiveYear Plan’’ Science and Technology Research Projects of Jilin Province (No. 2012-95) and the Jilin City Planning Project of Science and Technology (201132402) for their financial supports. References Antoniou, P., Hanilton, J., Koopman, B., Jain, R., Holloway, B., Lyberatos, G., Sboronos, S.A., 1990. Effect of temperature and pH on the effective maximum specific growth rate of nitrifying bacteria. Water Res. 24, 97–101. APHA, AWWA, WPCP, 2005. Standard Methods for the Examination of Water and Wastewater. In: Clesceri, L.S., Greenberg, A.E., Eaton, A.D. (Eds.), 21st ed. 21st ed. American Public Health Association, Washington, DC, USA. Baeza, J.A., Gabriel, D., Lafuente, J., 2002. Improving the nitrogen removal efficiency of an A2/O based WWTP by using an on-line knowledge based expert system. Water Res. 36, 2109–2123. Chiemchaisri, C., Yamamoto, K., 1993. Biological nitrogen removal under low temperature in a membrane separation bioreactor. Water Sci. Technol. 28, 325– 333. Chiu, Y.C., Lee, L.L., Chang, C.N., Chao, A.C., 2007. Control of carbon and ammonium ratio for simultaneous nitrification and denitrification in a sequencing batch bioreactor. Int. Biodeterior. Biodegrad. 59, 1–7. Choi, E., Rhu, D., Yun, Z., Lee, E., 1998. Temperature effects on biological nutrient removal system with weak municipal wastewater. Water Sci. Technol. 37, 219– 226. Ding, D., Feng, C., Jin, Y., Hao, C., Zhao, Y., Suemura, T., 2011. Domestic sewage treatment in a sequencing batch biofilm reactor (SBBR) with an intelligent controlling system. Desalination 276, 260–265. Fontenot, Q., Bonvillain, C., Kilgen, M., Boopathy, R., 2007. Effects of temperature, salinity, and carbon: nitrogen ratio on sequencing batch reactor treating shrimp aquaculture wastewater. Bioresour. Technol. 98, 1700–1703. He, S., Xue, G., Wang, B., 2009. Factors affecting simultaneous nitrification and denitrification (SND) and its kinetics model in membrane bioreactor. J. Hazard. Mater. 168, 704–710. Hocaoglu, S.M., Insel, G., Cokgor, E.U., Orhon, D., 2011a. Effect of low dissolved oxygen on simultaneous nitrification and denitrification in a membrane bioreactor treating black water. Bioresour. Technol. 102, 4333–4340. Hocaoglu, S.M., Insel, G., Cokgor, E.U., Orhon, D., 2011b. Effect of sludge age on simultaneous nitrification and denitrification in membrane bioreactor. Bioresour. Technol. 102, 6665–6672. Holakoo, L., Nakhla, G., Bassi, A.S., Yanful, E.K., 2007. Long term performance of MBR for biological nitrogen removal from synthetic municipal wastewater. Chemosphere 66, 849–857.
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