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anoxic denitrifying phosphorus removal process: The effects of carbon sources shock
Chemical Engineering Journal 172 (2011) 999–1007
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Chemical Engineering Journal journal homepage: www.elsevier.com/locate/cej
N2 O production in anaerobic/anoxic denitrifying phosphorus removal process: The effects of carbon sources shock Yayi Wang ∗ , Junjun Geng, Gang Guo, Chong Wang, Shanhu Liu State Key Laboratory of Pollution Control and Resources Reuse, School of Environmental Science and Engineering, Tongji University, Siping Road, Shanghai 200092, PR China
a r t i c l e
i n f o
Article history: Received 7 May 2011 Received in revised form 13 July 2011 Accepted 13 July 2011 Keywords: Nitrous oxide (N2 O) Denitrifying phosphorus removal Carbon sources Enzyme activity Microbial activity Oxidative stress
a b s t r a c t Nitrous oxide (N2 O) emissions from biological denitrification processes are one of the major sources of N2 O emissions during the biological nutrient removal process in wastewater treatment. In this study, the effects of carbon source shocks on N2 O production during denitrifying phosphorus removal were investigated using biomass that was initially acclimated with acetate. The shock tests were conducted in three laboratory-scale anaerobic/anoxic (An/A) reactors using acetate, acetate/propionate or propionate as the carbon sources. After switching the carbon source from acetate to acetate/propionate and propionate, the ratios of N2 O–N production to total nitrogen (TN) removal increased by 1.72 or 0.77 times, respectively, and the total nitrogen (TN) removal efficiency decreased from 80% to 52% or 38%, respectively; also, the PO4 3− –P removal efficiency declined from 75% to 63% or 47%, respectively. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Eutrophication is a major water quality problem worldwide that causes algal blooms and disrupts the normal functioning of aquatic ecosystems. Removal of nitrogen and phosphorus from sewage and industrial wastewater is one of the key strategies for preventing eutrophication. Denitrifying phosphorus removal in activated sludge systems has become an attractive alternative to traditional biological phosphorus removal because it enables simultaneous nitrogen and phosphorus removal via denitrifying phosphorus accumulating organisms (DPAOs), which are capable of using nitrate and/or nitrite as an electron acceptor for P removal instead of oxygen. This process reduces the demand for oxygen and carbon sources, which leads to reduced plant operational costs [1]. Nitrous oxide (N2 O) is an extremely potent greenhouse gas that is approximately 300 times more powerful than carbon dioxide (CO2 ) over its 120-year lifetime in the atmosphere [2]. N2 O is expected to be the dominant ozone-depleting substance emitted in the 21st century [3]. Numerous studies have reported that both nitrification and denitrification could lead to N2 O production in wastewater treatment systems [4,5]. A recent review of N2 O emissions from full-scale and lab-scale wastewater systems showed that they reached 0–14.6% of the nitrogen load in fullscale wastewater systems, whereas they ranged from 0 to 95% of
∗ Corresponding author. Tel.: +86 21 65984275; fax: +86 21 65984275. E-mail address: [email protected] (Y. Wang). 1385-8947/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.cej.2011.07.014
the nitrogen load in lab-scale wastewater systems [6]. However, in some simultaneous nitrogen and phosphorus removal systems, N2 O rather than N2 was found to be the major denitrification endproduct [7–9]. In addition, large amounts of N2 O have been found to be emitted from denitrifying phosphorus removal systems during the denitrification and phosphorus uptake period [10–12]. If these findings are correct, the potential advantages of denitrifying phosphorus removal techniques may not be as great as previously believed [12]. Different carbon sources, volatile fatty acids (VFA) and non-VFA have been shown to have important effects on the nutrient removal efficiency of biological nutrient removal (BNR) systems. Numerous studies have been conducted to characterize the effects of carbon sources on biological N and P removal [13–18]. However, only a few studies have focused on the impact of carbon sources on N2 O generation [19–21]. Li et al. [19] reported that acetate was a better carbon source for the promotion of denitrification efficiency and reduction of N2 O production than glucose or sucrose. Adouani et al. [20] examined the effects of carbon sources (acetate, ethanol, a mixture composed of ethanol and acetate, and two long carbon chain compounds) on N2 O emissions during biological denitrification. The results of their study revealed that the highest and lowest N2 O emissions occurred when acetate and ethanol were used as a carbon source, respectively. Later, Zhu and Chen [21] added sludge alkaline fermentation liquid to an anaerobic–aerobic (low dissolved oxygen (DO)) biological wastewater treatment process as a carbon source and demonstrated that N2 O generation was reduced by 68.7% when compared with that obtained when acetic acid was used as the sole
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carbon source. However, no studies have been conducted to investigate the effects of carbon source on N2 O production during the denitrifying phosphorus removal process. Acetic and propionic acid are known to be the two most common volatile fatty acids (VFAs) in municipal wastewater, and the ratio of acetic to propionic acid varies both diurnally and seasonally [14]. In denitrifying phosphorus removal systems, the composition of poly--hydroxyalkanoates (PHA) (including poly--hydroxybutyrate (PHB), poly--hydroxyvalerate (PHV) and poly-3-hydroxy-2-methylvalerate (PH2MV)) produced by phosphorus accumulating organisms (PAOs/DPAOs) during the anaerobic period varies based on the carbon source. However, PAOs/DPAOs are unable to metabolize these different types of PHA at the same rate in the subsequent anoxic phase [15]; therefore, the use of different carbon sources may lead to different properties of P metabolism and N2 O emission. Moreover, carbon sources change during practical wastewater treatment processes, often in a very short period of time. Such undesired fluctuations in influent can adversely affect PAOs/DPAOs within the activated sludge process, including their ability to metabolize N and P, which could lead to increased N2 O production. Little is known about the effects of shock carbon source on N2 O production in DPAOs dominant systems. In this study, the effects of a sudden change in carbon sources on N2 O production were investigated using three lab-scale anaerobic/anoxic (A/An) batch reactors. The goal of this study was to reveal the characteristics of N2 O production during denitrification phosphorus removal using different carbon sources, as well as the relationship between carbon source and anaerobic VFAs uptake and the amount and composition of PHA synthesized. The treatment performance of reactors facing shock loading of carbon sources are also discussed to provide a process control strategy to minimize N2 O production while maintaining effective P and N removal.
time controllers. The temperature was maintained at 25–30 ◦ C and the airflow rate and stirring speed were 40 L/h and 80 rpm, respectively. Sludge was obtained from a sewage treatment plant (Quyang Sewage Treatment Plant, Shanghai, China) that employed a biological nutrient removal process. One standard SBR cycle consisted of a rapid feeding (15 min), an anaerobic phase (90 min), an anoxic phase (210 min), an aerobic reaction (30 min), a settling phase (45 min), an effluent decanting phase (15 min) and a 75min idle phase. Therefore, the SBR operated with three 8 h cycles per day. The SBR was fed with synthetic wastewater. During the first 15 min feeding period, 5.5 L of synthetic wastewater were pumped into the reactor, while 100 mL of KNO3 solution were pulse added into the reactor at the beginning of the anoxic period, giving an initial NO3 − –N concentration of 35 mg/L. During steady-state operation, 125 mL of sludge was removed at the end of the aerobic period, giving a solids retention time (SRT) of approximately 20 days and a mixed liquor suspended solids (MLSS) and mixed liquor volatile suspended solids (MLVSS) level of about 3.7 and 2.3 g/L, respectively. The hydraulic retention time (HRT) of the SBR was approximately 10.9 h. The An/A/O SBR was operated over four months.
2. Materials and methods
Batch experiments were conducted to study the effects of carbon source shock on N2 O production during denitrifying phosphorus removal. These experiments were conducted after 80 days of the An/A/O SBR operation, when the SBR was working with a denitrifying phosphorus removal population in steady state. Three 2.5 L sealed reactors with a working volume of 2.4 L and an overhead space of 0.1 L were used for these experiments. The sludge (2 L) for the batch tests was obtained from the An/A/O SBR at the end of the decanting phase. After being washed with NaCl solution (0.154 M)
2.1. Long-term experiments in the anaerobic/anoxic/oxic sequencing batch reactor (An/A/O SBR) A sealed laboratory-scale An/A/O SBR with a working volume of 7.5 L and an overhead space of 1 L was used for the experiments (Fig. 1). The SBR was completely automated, and all peristaltic pumps, stirrers, air pumps and phase lengths were regulated by
2.2. Synthetic wastewater The synthetic wastewater used in this study contained (per liter): 448.46 mg of CH3 COONa (350 mg of chemical oxygen demand (COD)); 32.9 mg of KH2 PO4 (7.5 mg of P); 42 mg of K2 HPO4 (7.5 mg of P); 57.4 mg of NH4 Cl; 85 mg of MgSO4 ·7H2 O; 10 mg of CaCl2 ; and 110 mg of NaHCO3 . Additionally, the synthetic wastewater also contained the trace salt solution (0.3 mL/L) described by Wang et al. [18]. 2.3. Batch experiments
Fig. 1. The schematic diagram of experimental An/A/O SBR.
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three times, the sludge was evenly divided into three parts and then transferred into three batch reactors that were operated with a 90 min anaerobic reaction and 210 min anoxic reaction. Next, 1.7 L of synthetic wastewater were rapidly added to each reactor at the beginning of the cycle, which resulted in a MLVSS level of about 2.1 g/L in each bioreactor. The synthetic wastewater was similar to that of the An/A/O SBR, except for the carbon source. Three batch reactors fed with three different carbon sources, i.e. 350 mg COD of acetate, acetate/propionate (175 mg/L/175 mg/L) and propionate, were defined as R-A, R-M and R-P, respectively. At the end of the anaerobic period, KNO3 solution was pulse added into each reactor, giving an initial NO3 − –N concentration of 35 mg/L. The temperature for all tests was controlled at 28 ± 2 ◦ C. The pH values in the experiments were maintained at 7.5 ± 0.1 by adding 0.3 M HCl or 0.3 M NaOH. Batch experiments were conducted in duplicate at different An/A/O SBR operational times, and their average results and standard deviations were reported. The liquid and solid-phase samples were taken for chemical analysis with sampling intervals of 10 min during the first 30 min of the anaerobic period and anoxic period and each 30 min thereafter. Gas samples were collected using gas-tight collecting bags and syringes and immediately analyzed. 2.4. Analytical methods The liquid samples were immediately filtered through Millipore filter units (0.45 m pore size) for analysis of COD, NH4 + –N, NO3 − –N, NO2 − –N and PO4 3− –P. NH4 + –N, NO3 − –N, NO2 − –N, PO4 3− –P, MLSS and MLVSS were measured according to the Chinese State Environmental Protection Agency (SEPA) Standard Methods [22]. Merck COD reagents were used for the COD test (Merck; Darmstadt, Germany). The dissolved oxygen (DO), oxidation–reduction potential (ORP) and pH were measured online using oxygen, ORP and pH meters, respectively (Oxi 3310 and pH 3310, WTW Company, Germany). Total nitrogen (TN) was calculated as the sum of NH4 + –N, NO3 − –N and NO2 − –N. Free nitrous acid nitrogen (FNA-N) was calculated according to the method described by Anthonisen et al. [23]. Glycogen was determined using the method described by Jenkins et al. [24]. The transformation rates of NO3 − –N, PO4 3− –P, and VFAs were determined through linear regression of the measured profiles, and the reduction rate of NO2 − –N was determined through linear regression of the measured NOx − –N (NO3 − –N + NO2 − –N) profiles. The approximate reduction rate of N2 O–N was determined by linear regression of the measured NO3 − –N + NO2 − –N + N2 O–N profiles. The N2 O concentrations in both gas and liquid samples were analyzed using a gas chromatograph (GC) (Agilent 7820, USA) [12]. Acetic acid (HAc) and propionic acid (Pro) were measured according to the method described by Wang et al. [12]. PHB, PHV and PH2MV were measured according to the method described by Oehmen et al. [25]. The total PHA in the samples was calculated as the sum of the measured PHB, PHV and PH2MV. Dehydrogenase activity was determined according to the method described by Goel et al. [26], with minor modification. 3. Results and discussion 3.1. Cyclic studies in An/A/O SBR with acetate as a sole carbon source Cyclic studies were conducted in the An/A/O SBR to establish steady state operation and determine the reactor phenotypes. Fig. 2 shows a typical example of the concentration profiles measured during an SBR cycle after stable N and P conversions were achieved. The An/A/O SBR displayed a typical DPAOs phenotype.
Fig. 2. Variations in N, P, N2 O, HAc, glycogen and PHA in one cycle in the A/An/O SBR.
During the anaerobic phase, the acetate was typically completely consumed within 60 min, and this was accompanied by the formation of PHA, release of phosphorus and consumption of glycogen. In the subsequent anoxic period, once KNO3 solution was added to the reactor, simultaneous phosphorus uptake and denitrification occurred rapidly, concurrently with PHA degradation and glycogen replenishment. NO3 − –N was depleted after 60 min of anoxic operation. The NO2 − –N increased to 20 mg/L at 120 min, after which it was gradually reduced. Overall, approximately 2.60 mg N/L of N2 O was produced during the anoxic phase, giving a ratio of N2 O–N production to TN removal of 7.77% (Table 1). Regarding the post oxic phase, a small amount of NH4 + –N was oxidized to NO3 − –N or Table 1 Comparison of N and P removal and N2 O production in one cycle in the A/An/O SBR. Item
Values
MLSS (mg/L) MLVSS (mg/L) PO4 3− –P removal efficiency at the end of anoxic period (%)a PO4 3− –P removal efficiency at the end of oxic period (%)b TN removal efficiency (%)c Anoxic N2 O production (mg N/L) Oxic N2 O production (mg N/L) Ratio of anoxic N2 O–N production to TN removal (%)
3483 2002 82% 89% 77% 2.60 0.36 7.77%
a Anoxic end PO4 3− –P removal efficiency (%) = (P0 − P1 )/P0 ; P0 : PO4 3− –P concentration at the beginning of the anaerobic phase; P1 : PO4 3− –P concentration at the end of the anoxic phase. b Oxic end PO4 3− –P removal efficiency (%) = (P0 − P2 )/P0 ; P0 : P concentration at the beginning of the anaerobic phase; P2 : PO4 3− –P concentration at the end of the aerobic stage. c TN removal efficiency (%) = (TN1 − TN2 )/TN1 ; TN1 : TN concentration at the beginning of the anoxic phase; TN2 : TN concentration at the end of the anoxic phase.
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Table 2 Anaerobic and anoxic transformations of PHA and glycogen in one cycle in the A/An/O SBR. Item
Values
PHB synthesis (mmol-C/g-MLVSS) PHV synthesis (mmol-C/g-MLVSS) PHA synthesis (mmol-C/g-MLVSS) PHA synthesis/Hac uptake (mmol-C/mmol-C) Phosphorus release/Hac uptake (mmol-P/mmol-C) Glycogen degradation (mmol-C/g-MLVSS) PHB degradation (mmol-C/g-MLVSS) PHV degradation (mmol-C/g-MLVSS) PHA degradation (mmol-C/g-MLVSS) Glycogen synthesis (mmol-C/g-MLVSS) PHB synthesis/PHA synthesis (mmol-C/mmol-C) PHV synthesis/PHA synthesis (mmol-C/mmol-C) Hac uptake rate (mmol-C/g-MLVSS·h)
2.43 0.98 3.41 0.86 0.07 2.15 2.08 0.67 2.77 1.70 71% 29% 3.94
NO2 − –N (Fig. 2a), and this was accompanied by PHA degradation, glycogen replenishment and phosphorus assimilation (Fig. 2b). Moreover, approximately 0.36 mg N2 O–N/L was produced during this stage (Table 1), possibly via the denitrification pathway [12,27]. Kim et al. [27] reported that N2 O was generated in the aerobic nitrification stage when NO2 − –N and NH4 + –N existed simultaneously, and they deemed that N2 O was produced via the denitrification pathway by ammonia-oxidizing bacteria (AOB). It should be noted that the ratio of P release to Hac uptake for the An/A/O SBR was only 0.07 mmol-P/mmol-C (Table 2), which was significantly smaller than the ratios of 0.2–0.5 mmol-P/mmolC reported in previous studies [12,14,15,28]. Moreover, a high amount of PHV (29% of PHA) was produced during the anaerobic phase (Table 2). Taken together, these findings appear to indicate that glycogen accumulating organisms (GAOs) were present and active in the system and contributed to carbon source storage. GAOs can take up VFAs to store PHA under anaerobic conditions
by using glycogen as the intracellular energy pool instead of poly-P and not accumulating polyphosphate under aerobic/anoxic conditions, which results in competition with PAOs. Coincidently, GAOs tend to produce more PHV than PAOs via the propionate–succinate pathway [29]. Based on the long-term operation of the An/A/O SBR, the average effluent PO4 3− –P was high (3.49 mg/L) and the average PO4 3− –P removal efficiency was only 77% (data not shown), which confirmed that GAOs had survived in the SBR. The enrichment of GAOs in the studied SBR was primarily because of the high temperature (25–30 ◦ C), as temperatures over 20 ◦ C favor the growth of GAOs [30].
3.2. Shock effects of different carbon sources on denitrifying phosphorus removal and N2 O production in three batch reactors Batch tests were conducted with acetate-enriched biomass in the An/A/O SBR using acetate, mixed acetate and propionate, and propionate alone as the carbon sources.
3.2.1. Comparison of VFAs uptake in R-A, R-M and R-P As shown in Fig. 3f, the VFAs were completely taken up in R-A and R-M, while a fraction of propionic acid still remained in R-P at the end of the anaerobic stage. Additionally, the Hac uptake rate was higher in R-A than in R-M, and the propionic acid uptake rate increased by 2.95 times in R-M when compared to R-P (Table 3). Moreover, the propionic acid uptake rate was 1.26 times that of the Hac uptake rate in the R-M system. These findings suggest that the presence of Hac stimulated propionic acid uptake, while the presence of propionic acid inhibited Hac uptake. Similar results were also obtained by Li et al. [16], who examined the long-term effects of carbon source in an anaerobic–aerobic (low DO) phosphorus and nitrogen removal system and found that both propionic acid and
Fig. 3. Variations in N, P, N2 O and VFAs in R-A, R-M and R-P.
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Table 3 Anaerobic and anoxic transformations of PHA and glycogen in batch experiments. Item
Hac (R-A)
PHA synthesis/VFA uptake (mmol-C/mmol-C) Anaerobic end PHB (mmol-C/g-MLVSS) Anaerobic end PHV (mmol-C/g-MLVSS) Anaerobic end PHA (mg COD/L) PHB synthesis/PHA synthesis (mmol-C/mmol-C) PHV synthesis/PHA synthesis (mmol-C/mmol-C) Hac uptake rate (mmol-C/g-MLVSS h) Pro uptake rate (mmol-C/g-MLVSS h)
1.01 2.90 1.18 311 74.21% 25.66% 6.30 –
± ± ± ± ± ± ±
Hac uptake rates decreased as the ratio of propionic to acetic acid increased. Notably, the propionic acid uptake rates in R-M and R-P were 72% and 24% of the Hac uptake rate in R-A, respectively (Table 3). These findings indicate that biomass acclimatized with acetate may not be able to anaerobically take up propionic acid as quickly as Hac after only one cycle. Similarly, Oehmen et al. [15] found that the propionic acid uptake rate declined after a switch in carbon source from Hac to propionic acid, and that it decreased more as the fraction of GAOs increased. 3.2.2. Effects of carbon sources on the transformation of PHA and glycogen Fig. 4 shows the transformations of PHA and glycogen in R-A, R-M and R-P. When the carbon source was switched from acetate (R-A) to acetate/propionate (R-M) and then to propionate (R-P), the PHA synthesis decreased from 3.59 to 2.01 and then to 1.11 mmolC/g-MLVSS (Fig. 5a). Moreover, a decrease and an increase occurred
Glycogen (mol-C/g-MLVSS)
9
anoxic
anae robic
8
a)
7 6 Hac Hac+Pro Pro
5 4
PHA, PH2MV (mmol-C/g-MLVSS)
PHB, PHV (mmol-C/g-MLVSS)
3 5
b)
4
PHB (Hac) PHV (Hac) PHB(Hac+Pro) PHV (Hac+P ro) PHB (Pro) PHV (Pro)
3 2 1 0 5
c)
4
PHA (Hac) PH2MV (Hac) PHA (Hac+P ro) PH2MV (Hac+Pro) PHA (Pro) PH2MV (Pro)
3 2 1 0 0
30
60
90
12 0
15 0
18 0
210
24 0
270
time (min) Fig. 4. Variations of glycogen and PHA in R-A, R-M and R-P.
30 0
Hac + Pro (R-M) 0.03 0.08 0.16 19 1.71% 1.85% 0.24
0.61 0.46 2.06 203 14.59% 85.31% 3.60 4.55
± ± ± ± ± ± ± ±
0.01 0.15 0.23 30 0.59% 0.85% 0.01 0.23
Pro (R-P) 0.44 0.30 1.39 140 9.92% 89.61% – 1.54
± ± ± ± ± ±
0.06 0.07 0.20 22 2.02% 1.54%
± 0.33
in the fractions of formed PHB and PHV, respectively, as the percentage of added propionate increased (Table 3). These findings corresponded with those of Li et al. [16], who investigated an anaerobic/aerobic (low DO) process. Moreover, according to Lemos et al. [31], the primary fractions of PHA produced by PAOs were PHV and PH2MV when the carbon source was propionate. However, in the R-M and R-P systems, negligible amounts of PH2MV were anaerobically synthesized. Pijuan et al. [32] reported a similar finding, in which PH2MV was not synthesized when the carbon source was switched from acetate to propionate. However, further research is needed to clarify this issue. The ratios of PHA synthesis to VFA uptake (mmol-C/mmol-C) were lower in R-M and R-P (Table 3). This finding is in accordance with the results of several other studies [16,32] that revealed that propionate addition led to lower PHA synthesis/VFA uptake. It is also possible that some forms of PHA were not quantified, e.g. poly3-hydroxy-2-methylbutyrate (PH2MB). Lemos et al. [31] showed that synthesized PH2MB comprised 8.6% of the PHA formed when propionic acid was used as a carbon source by using 13 C-labeling and in vivo nuclear magnetic resonance. Moreover, it should be noted that about 0.6 mmol-C/g-MLVSS of PHA was degraded during the latter part of the anaerobic phase in R-M (Fig. 4b and c), leading to a lower net PHA synthesis/VFA uptake in R-M. Similar findings have been reported in our previous studies [12]. The amount of anaerobic glycogen degradation decreased when the carbon source was changed from acetate (R-A) to mixed acetate and propionate (R-M) and propionate (R-P), which corresponded well with the trend of the amount of PHA synthesis in each reactor (Fig. 5a and c). These findings were supported by the observations of Chen et al. [14] and Li et al. [16], who found that the increase in the influent ratio of propionic to acetic acid resulted in a decrease in glycogen degradation. During the anoxic phase, the amounts of PHA degradation in R-A, R-M and R-P were 3.94, 2.42 and 1.45 mmol-C/g-MLVSS, respectively (Fig. 5b), suggesting that increasing propionic acid content results in decreased PHA oxidation. In addition, the glycogen transformation showed a trend similar to the PHA degradation (Fig. 5b and c). 3.2.3. Comparison of N and P removal performance in R-A, R-M and R-P Of R-A, R-M, and R-P, which utilized acetate, acetate/propionate and propionate as the carbon source, respectively, the highest and lowest P release were observed in R-A and R-M, respectively (Figs. 3e and 5d). Moreover, anaerobic P uptake occurred in R-A (about 5.0 mg/L) and R-M (about 10.6 mg/L) during the latter part of the anaerobic phase, leading directly to a reduced net P release [12]. Therefore, the actual ability of P release was highest in R-A and lowest in R-P when acetate and propionate were used as the sole carbon source. This coincides well with other studies that have suggested that an increase in the propionic/acetic ratio led to less P release [16]. After the three reactors were fed with KNO3 solution, denitrification and P uptake occurred simultaneously. The highest P
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Y. Wang et al. / Chemical Engineering Journal 172 (2011) 999–1007
4.0
5.0
a
b
PHB synthesis PHV synthesis PHA synthesis
3.5
4.0 3.5 mmol-C/g-MLVSS
mmol-C/g-MLVSS
3.0 2.5 2.0 1.5 1.0
3.0 2.5 2.0 1.5 1.0
0.5
0.5 0.0
0.0 Hac (R-A)
3.0
Hac+Pro (R-M)
Hac (R-A)
Pro (R-P)
60
c
glycogen degradation glycogen syhtnesis
50
2.0
1.5
1.0
0.5
d
Hac+Pro (R-M)
phosphorus release phosphorus uptake TN removal
Pro (R-P)
100 phosphorus removal efficency TN removal efficency
90 80 70
40
60
30
50 40
20
30
removal efficency (%)
nitrogen or phosphorus (mg/L)
2.5 mmol-C/g-MLVSS
PHB degradation PHV degradation PHA degradation
4.5
20
10
10
0
0.0 Hac (R-A)
Hac+Pro (R-M)
Pro (R-P)
0
R-A (Hac)
R-M (Hac+Pro)
R-P (Pro)
Fig. 5. The effects of carbon sources on N and P removal and the transformations of glycogen and PHA in R-A, R-M and R-P.
removal was obtained in R-A, while the lowest was observed in R-P (Fig. 5d). This is in good agreement with the observation that a higher propionic acid content led to detrimental P removal in the short term in an anaerobic–oxic SBR [14]. Moreover, the added NO3 − –N was rapidly reduced in R-A, R-M and R-P, and this reduction was accompanied by the accumulation and reduction of NO2 − –N (Fig. 3b and c), which resulted in corresponding TN removal efficiencies of 80%, 52% and 38%, respectively. These findings indicate that, as the amount of propionate added increased, the denitrifying phosphorus removal decreased, based on only one cycle. 3.2.4. Effects of carbon sources on N2 O production in R-A, R-M and R-P Fig. 3d shows the variations in off-gas, dissolved and total N2 O concentrations in R-A, R-M and R-P. During the anaerobic phase, no N2 O production was detected in any of the reactors (Fig. 3d). However, after the KNO3 solution was pulse added at the end of the anaerobic period, a remarkable increase in the N2 O concentration occurred in all reactors. Thereafter, the N2 O production rate decreased gradually, and this was followed by a decrease in the N2 O concentration (Fig. 3d). Consequently, the ratios of N2 O–N production to the TN removal were 13.93 ± 0.05%, 37.95 ± 5.62% and 24.71 ± 4.59%, respectively, in R-A, R-M and R-P (Table 4). Obviously, when the carbon source was switched from acetate (R-A) to acetate/propionate (R-M) and propionate (R-P), the ratio of N2 O–N production to TN removal increased by 1.72 and 0.77 times, respectively. These findings suggest that changing the carbon source could lead to more N2 O production in one cycle during the denitrify-
ing phosphorus removal process. Moreover, the dynamic changes in carbon source types could produce a shock loading effect on the biomass in wastewater treatment bioreactors, which tends to result in poor nutrient removal efficiencies and a large amount of N2 O production. These scenarios also occur in practical wastewater treatment processes. 3.3. Factors leading to increased N2 O production after a change in carbon sources 3.3.1. Effect of activities of denitrifying enzymes The carbon source may affect the activities of denitrifying enzymes, leading to a different amount of N2 O production [13,17,19,21]. As shown in Table 4, when the carbon source was switched from acetate to acetate/propionate or propionate, the rates of NO3 − –N and NO2 − –N reduction both decreased, suggesting that the activities of nitrate and nitrite reductase were decreased. Furthermore, the carbon source shock can produce different effects on nitrate and nitrite reductase, which led to different ratios of the NO3 − –N reduction rate to the NO2 − –N reduction rate in R-A, RM and R-P (Table 4). Moreover, the NO3 − –N reduction rate was much higher than the NO2 − –N reduction rate (Table 4), and this imbalance caused a serious accumulation of NO2 − –N throughout the anoxic stage in the three reactors. This NO2 − –N accumulation increased as the ratio of the NO3 − –N reduction rate to the NO2 − –N reduction rate increased (Table 4 and Fig. 3c). Nitrite and FNA are known to inhibit the activity of nitrous oxide reductase (Nos), which causes the accumulation of N2 O in denitrification processes [10,11]. As a result, the greatest and lowest ratios of N2 O–N
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Table 4 Comparison of N and P removal and N2 O production in batch experiments. Item
Hac (R-A)
Hac + Pro (R-M)
Pro (R-P)
MLSS (mg/L) MLVSS (mg/L) NO3 − –N reduction rate (mg-N/g-MLVSS h) NO2 − –N reduction rate (mg-N/g-MLVSS h) NO3 − –N reduction rate/NO2 − –N reduction rate N2 O–N reduction rate (mg-N/g-MLVSS h) Dehydrogenase activity (mg TFa /g MLVSS h) PO4 3− –P uptake rate (mg-P/g-MLVSS h) Anoxic N2 O production (mg N/L) Ratio of anoxic N2 O–N production to TN removal (%)
3578 ± 49 2075 ± 68 16.60 ± 0.67 7.38 ± 0.95 2.25 5.73 ± 1.10 26.04 ± 2.53 19.82 ± 1.52 4.87 ± 0.13 13.93 ± 0.05
3619 ± 42 2117 ± 28 16.00 ± 0.76 5.56 ± 0.90 2.88 2.16 ± 0.32 21.08 ± 0.90 12.53 ± 0.40 7.57 ± 1.23 37.95 ± 5.62
3741 ± 62 2180 ± 79 11.71 ± 0.89 4.20 ± 0.37 2.79 1.68 ± 0.26 19.20 ± 1.53 8.18 ± 0.23 4.17 ± 1.27 24.71 ± 4.59
a
TF, triphenylformazan.
production to TN removal were observed in R-M and R-A, respectively, as the highest and lowest ratios of the NO3 − –N reduction rate to the NO2 − –N reduction rate occurred in R-M (2.88) and R-A (2.25), respectively (Table 4). Notably, the N2 O reduction rate in R-M and R-P was much lower than that in R-A (Table 4), suggesting that the activities of Nos were lower in R-M and R-P than in R-A. This further explained why the amount of N2 O produced by R-M and R-P was higher than that produced by R-A.
3.3.2. Effects of the amount and composition of anaerobically synthesized PHA Carbon source concentration is an important factor influencing N2 O production under anoxic conditions [33]. In our study, the carbon source for denitrification was from anaerobically synthesized PHA. Therefore, the amount of anaerobically synthesized PHA played an important role in N2 O production [12]. As shown in Table 3 and Fig. 5a, the levels of anaerobically synthesized PHA in R-A, R-M and R-P were 3.59, 2.01 and 1.11 mmol-C/g-MLVSS, corresponding to 311, 203 and 140 mg COD/L of PHA at the end of anaerobic stage, respectively. Obviously, the lower COD availability in R-M and R-P during the initial anoxic phase led to more N2 O production when compared to R-A. Additionally, even though more COD was available in R-M than in R-P, more N2 O was produced in R-M than in R-P. A possible explanation for this is that the propionate that was not consumed during the anaerobic period was residual to the anoxic phase in R-P, during which it was exhausted by ordinary denitrifying bacteria, which improved the denitrification rate and reduced the N2 O emissions [12]. Nevertheless, more NO2 − –N was accumulated in R-M than in R-P (Fig. 3c), which also attributed to the higher N2 O production in R-M than in R-P. Carbon source types have been reported to have impacts on N2 O emissions during biological denitrification [19–21]. In the present study, an internal carbon source (PHA) was used for biological denitrification, and the different composition of PHA (mainly PHB and PHV in the present study) may also be an important factor influencing N2 O production. As shown in Fig. 5a, the highest and lowest amount of PHV was formed in R-M and R-A, respectively, which corresponded to the greatest and lowest ratio of N2 O–N production to TN removal. These findings suggest that more PHV leads to higher N2 O production. As shown in Fig. 6, the metabolic pathway involved conversions of PHV in DPAOs that differed from those of PHB, as reported by Martin et al. [34]. Consequently, the organisms (PAOs, DPAOs or GAOs) may not be able to anoxically metabolize PHV as quickly as PHB in a transition period of only one cycle [15], which subsequently reduced the activities of the PAOs/DPAOs or GAOs. Moreover, the dehydrogenase activity, which reflects the microbial activity [26], declined after a switch in carbon source from acetate to acetate/propionate or propionate (Table 4). Similarly, the PO4 3− –P uptake rate and the NO3 − –N and NO2 − –N reduction rates
all decreased when the influent carbon source was switched from acetate (R-A) to mixed substrates (R-M) and then to propionate (RP) (Table 4). Taking into account all of these factors, it is clear that the use of PHV as a carbon source could result in a decrease in PAOs or GAOs activity after a transition in carbon sources in the short term, thereby triggering increased N2 O production. However, it is difficult to determine if the lack of acclimation of the biomass to these substrates was the reason for the high N2 O emissions, or if they were the normal N2 O emissions for these substrates. Accordingly, the long-term effects of PHV on N2 O production require further study. 3.3.3. Effects of oxidative stress on N2 O production in R-A, R-M and R-P The induction of oxidative stress in biological systems has been reported under unfavorable environmental or stress conditions [35,36]. Oxidative stress may cause detrimental effects in many cell types, and an oxidative stress response in bacteria has been implicated as the cause of process failure in wastewater treatment facilities. Moreover, oxidative stress could induce activation of the antioxidant response element (ARE), leading to the expression of cytoprotective enzymes, such as superoxide dismutase (SOD), which was detected by Martin et al. [34] in EBPR sludge. Superoxide dismutase catalyzes the reduction of superoxide anions into H2 O2 [37].
Fig. 6. Schematic diagram of the anoxic metabolic pathway for conversion of PHB and PHV in DPAOs.
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Although no studies have confirmed that oxidative stress occurs when the biomass being treated is subjected to a sudden switch in carbon source, we speculate that oxidative stress conditions may occur in our studied R-M and R-P systems. As shown in Fig. 3b, when the carbon source was switched from acetate to acetate/propionate or propionate, the NO3 − –N concentrations increased during the latter part of the anoxic period in R-M and R-P, indicating that nitrification reactions occurred. Conversely, this phenomenon was not observed in R-A. Therefore, we assume that trace of O2 was present in R-M and R-P. In the batch tests, the reactors were properly sealed off from the outside atmosphere and the DO level during the entire anoxic period was lower than 0.1 mg/L. However, oxidative stress associated with reduction of H2 O2 can produce O2 [36,37]. Sabumon P C also observed the anoxic nitrification, and suggested that trace of O2 was produced from the reduction of H2 O2 [36]. These findings showed the occurrence of oxidative stress in R-M and R-P. Furthermore, oxidative stress has adverse effects on DPAOs, DGAOs and enzymes [35], leading directly to increased N2 O production in R-M and R-P when compared to R-A. 4. Conclusions Carbon source shock significantly influenced N2 O production during the denitrification phosphorus removal process. The N2 O–N production to TN removal ratios were 1.72 or 0.77 times higher, respectively, when biomass acclimatized with acetate was subjected to acetate/propionate or propionate shock than when acetate alone was used. The shock change of carbon sources significantly influenced the activities of denitrifying enzymes and the amount and composition of anaerobically synthesized PHA, and also caused oxidative stress. This directly stimulated N2 O production. In addition, the TN removal efficiency decreased from 80% to 52% or 38%, respectively, and the PO4 3− –P removal efficiency declined from 75% to 63% or 47%, respectively, after the carbon source was switched from acetate to acetate/propionate or propionate. Therefore, acetate/propionate or propionate was unfavorable for conducting denitrifying phosphorus removal in only one cycle.
Acknowledgments This study was supported by the National Natural Science Foundation of China (NSFC) (no. 51078283), the Program for New Century Excellent Talent in University (no. NCET-08-0404), the Shanghai Science and Technology Development Funds (no. 09QA1406100) and the Fundamental Research Funds for the Central University (Tongji University).
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Chemical Engineering Journal journal homepage: www.elsevier.com/locate/cej
N2 O production in anaerobic/anoxic denitrifying phosphorus removal process: The effects of carbon sources shock Yayi Wang ∗ , Junjun Geng, Gang Guo, Chong Wang, Shanhu Liu State Key Laboratory of Pollution Control and Resources Reuse, School of Environmental Science and Engineering, Tongji University, Siping Road, Shanghai 200092, PR China
a r t i c l e
i n f o
Article history: Received 7 May 2011 Received in revised form 13 July 2011 Accepted 13 July 2011 Keywords: Nitrous oxide (N2 O) Denitrifying phosphorus removal Carbon sources Enzyme activity Microbial activity Oxidative stress
a b s t r a c t Nitrous oxide (N2 O) emissions from biological denitrification processes are one of the major sources of N2 O emissions during the biological nutrient removal process in wastewater treatment. In this study, the effects of carbon source shocks on N2 O production during denitrifying phosphorus removal were investigated using biomass that was initially acclimated with acetate. The shock tests were conducted in three laboratory-scale anaerobic/anoxic (An/A) reactors using acetate, acetate/propionate or propionate as the carbon sources. After switching the carbon source from acetate to acetate/propionate and propionate, the ratios of N2 O–N production to total nitrogen (TN) removal increased by 1.72 or 0.77 times, respectively, and the total nitrogen (TN) removal efficiency decreased from 80% to 52% or 38%, respectively; also, the PO4 3− –P removal efficiency declined from 75% to 63% or 47%, respectively. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Eutrophication is a major water quality problem worldwide that causes algal blooms and disrupts the normal functioning of aquatic ecosystems. Removal of nitrogen and phosphorus from sewage and industrial wastewater is one of the key strategies for preventing eutrophication. Denitrifying phosphorus removal in activated sludge systems has become an attractive alternative to traditional biological phosphorus removal because it enables simultaneous nitrogen and phosphorus removal via denitrifying phosphorus accumulating organisms (DPAOs), which are capable of using nitrate and/or nitrite as an electron acceptor for P removal instead of oxygen. This process reduces the demand for oxygen and carbon sources, which leads to reduced plant operational costs [1]. Nitrous oxide (N2 O) is an extremely potent greenhouse gas that is approximately 300 times more powerful than carbon dioxide (CO2 ) over its 120-year lifetime in the atmosphere [2]. N2 O is expected to be the dominant ozone-depleting substance emitted in the 21st century [3]. Numerous studies have reported that both nitrification and denitrification could lead to N2 O production in wastewater treatment systems [4,5]. A recent review of N2 O emissions from full-scale and lab-scale wastewater systems showed that they reached 0–14.6% of the nitrogen load in fullscale wastewater systems, whereas they ranged from 0 to 95% of
∗ Corresponding author. Tel.: +86 21 65984275; fax: +86 21 65984275. E-mail address: [email protected] (Y. Wang). 1385-8947/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.cej.2011.07.014
the nitrogen load in lab-scale wastewater systems [6]. However, in some simultaneous nitrogen and phosphorus removal systems, N2 O rather than N2 was found to be the major denitrification endproduct [7–9]. In addition, large amounts of N2 O have been found to be emitted from denitrifying phosphorus removal systems during the denitrification and phosphorus uptake period [10–12]. If these findings are correct, the potential advantages of denitrifying phosphorus removal techniques may not be as great as previously believed [12]. Different carbon sources, volatile fatty acids (VFA) and non-VFA have been shown to have important effects on the nutrient removal efficiency of biological nutrient removal (BNR) systems. Numerous studies have been conducted to characterize the effects of carbon sources on biological N and P removal [13–18]. However, only a few studies have focused on the impact of carbon sources on N2 O generation [19–21]. Li et al. [19] reported that acetate was a better carbon source for the promotion of denitrification efficiency and reduction of N2 O production than glucose or sucrose. Adouani et al. [20] examined the effects of carbon sources (acetate, ethanol, a mixture composed of ethanol and acetate, and two long carbon chain compounds) on N2 O emissions during biological denitrification. The results of their study revealed that the highest and lowest N2 O emissions occurred when acetate and ethanol were used as a carbon source, respectively. Later, Zhu and Chen [21] added sludge alkaline fermentation liquid to an anaerobic–aerobic (low dissolved oxygen (DO)) biological wastewater treatment process as a carbon source and demonstrated that N2 O generation was reduced by 68.7% when compared with that obtained when acetic acid was used as the sole
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carbon source. However, no studies have been conducted to investigate the effects of carbon source on N2 O production during the denitrifying phosphorus removal process. Acetic and propionic acid are known to be the two most common volatile fatty acids (VFAs) in municipal wastewater, and the ratio of acetic to propionic acid varies both diurnally and seasonally [14]. In denitrifying phosphorus removal systems, the composition of poly--hydroxyalkanoates (PHA) (including poly--hydroxybutyrate (PHB), poly--hydroxyvalerate (PHV) and poly-3-hydroxy-2-methylvalerate (PH2MV)) produced by phosphorus accumulating organisms (PAOs/DPAOs) during the anaerobic period varies based on the carbon source. However, PAOs/DPAOs are unable to metabolize these different types of PHA at the same rate in the subsequent anoxic phase [15]; therefore, the use of different carbon sources may lead to different properties of P metabolism and N2 O emission. Moreover, carbon sources change during practical wastewater treatment processes, often in a very short period of time. Such undesired fluctuations in influent can adversely affect PAOs/DPAOs within the activated sludge process, including their ability to metabolize N and P, which could lead to increased N2 O production. Little is known about the effects of shock carbon source on N2 O production in DPAOs dominant systems. In this study, the effects of a sudden change in carbon sources on N2 O production were investigated using three lab-scale anaerobic/anoxic (A/An) batch reactors. The goal of this study was to reveal the characteristics of N2 O production during denitrification phosphorus removal using different carbon sources, as well as the relationship between carbon source and anaerobic VFAs uptake and the amount and composition of PHA synthesized. The treatment performance of reactors facing shock loading of carbon sources are also discussed to provide a process control strategy to minimize N2 O production while maintaining effective P and N removal.
time controllers. The temperature was maintained at 25–30 ◦ C and the airflow rate and stirring speed were 40 L/h and 80 rpm, respectively. Sludge was obtained from a sewage treatment plant (Quyang Sewage Treatment Plant, Shanghai, China) that employed a biological nutrient removal process. One standard SBR cycle consisted of a rapid feeding (15 min), an anaerobic phase (90 min), an anoxic phase (210 min), an aerobic reaction (30 min), a settling phase (45 min), an effluent decanting phase (15 min) and a 75min idle phase. Therefore, the SBR operated with three 8 h cycles per day. The SBR was fed with synthetic wastewater. During the first 15 min feeding period, 5.5 L of synthetic wastewater were pumped into the reactor, while 100 mL of KNO3 solution were pulse added into the reactor at the beginning of the anoxic period, giving an initial NO3 − –N concentration of 35 mg/L. During steady-state operation, 125 mL of sludge was removed at the end of the aerobic period, giving a solids retention time (SRT) of approximately 20 days and a mixed liquor suspended solids (MLSS) and mixed liquor volatile suspended solids (MLVSS) level of about 3.7 and 2.3 g/L, respectively. The hydraulic retention time (HRT) of the SBR was approximately 10.9 h. The An/A/O SBR was operated over four months.
2. Materials and methods
Batch experiments were conducted to study the effects of carbon source shock on N2 O production during denitrifying phosphorus removal. These experiments were conducted after 80 days of the An/A/O SBR operation, when the SBR was working with a denitrifying phosphorus removal population in steady state. Three 2.5 L sealed reactors with a working volume of 2.4 L and an overhead space of 0.1 L were used for these experiments. The sludge (2 L) for the batch tests was obtained from the An/A/O SBR at the end of the decanting phase. After being washed with NaCl solution (0.154 M)
2.1. Long-term experiments in the anaerobic/anoxic/oxic sequencing batch reactor (An/A/O SBR) A sealed laboratory-scale An/A/O SBR with a working volume of 7.5 L and an overhead space of 1 L was used for the experiments (Fig. 1). The SBR was completely automated, and all peristaltic pumps, stirrers, air pumps and phase lengths were regulated by
2.2. Synthetic wastewater The synthetic wastewater used in this study contained (per liter): 448.46 mg of CH3 COONa (350 mg of chemical oxygen demand (COD)); 32.9 mg of KH2 PO4 (7.5 mg of P); 42 mg of K2 HPO4 (7.5 mg of P); 57.4 mg of NH4 Cl; 85 mg of MgSO4 ·7H2 O; 10 mg of CaCl2 ; and 110 mg of NaHCO3 . Additionally, the synthetic wastewater also contained the trace salt solution (0.3 mL/L) described by Wang et al. [18]. 2.3. Batch experiments
Fig. 1. The schematic diagram of experimental An/A/O SBR.
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three times, the sludge was evenly divided into three parts and then transferred into three batch reactors that were operated with a 90 min anaerobic reaction and 210 min anoxic reaction. Next, 1.7 L of synthetic wastewater were rapidly added to each reactor at the beginning of the cycle, which resulted in a MLVSS level of about 2.1 g/L in each bioreactor. The synthetic wastewater was similar to that of the An/A/O SBR, except for the carbon source. Three batch reactors fed with three different carbon sources, i.e. 350 mg COD of acetate, acetate/propionate (175 mg/L/175 mg/L) and propionate, were defined as R-A, R-M and R-P, respectively. At the end of the anaerobic period, KNO3 solution was pulse added into each reactor, giving an initial NO3 − –N concentration of 35 mg/L. The temperature for all tests was controlled at 28 ± 2 ◦ C. The pH values in the experiments were maintained at 7.5 ± 0.1 by adding 0.3 M HCl or 0.3 M NaOH. Batch experiments were conducted in duplicate at different An/A/O SBR operational times, and their average results and standard deviations were reported. The liquid and solid-phase samples were taken for chemical analysis with sampling intervals of 10 min during the first 30 min of the anaerobic period and anoxic period and each 30 min thereafter. Gas samples were collected using gas-tight collecting bags and syringes and immediately analyzed. 2.4. Analytical methods The liquid samples were immediately filtered through Millipore filter units (0.45 m pore size) for analysis of COD, NH4 + –N, NO3 − –N, NO2 − –N and PO4 3− –P. NH4 + –N, NO3 − –N, NO2 − –N, PO4 3− –P, MLSS and MLVSS were measured according to the Chinese State Environmental Protection Agency (SEPA) Standard Methods [22]. Merck COD reagents were used for the COD test (Merck; Darmstadt, Germany). The dissolved oxygen (DO), oxidation–reduction potential (ORP) and pH were measured online using oxygen, ORP and pH meters, respectively (Oxi 3310 and pH 3310, WTW Company, Germany). Total nitrogen (TN) was calculated as the sum of NH4 + –N, NO3 − –N and NO2 − –N. Free nitrous acid nitrogen (FNA-N) was calculated according to the method described by Anthonisen et al. [23]. Glycogen was determined using the method described by Jenkins et al. [24]. The transformation rates of NO3 − –N, PO4 3− –P, and VFAs were determined through linear regression of the measured profiles, and the reduction rate of NO2 − –N was determined through linear regression of the measured NOx − –N (NO3 − –N + NO2 − –N) profiles. The approximate reduction rate of N2 O–N was determined by linear regression of the measured NO3 − –N + NO2 − –N + N2 O–N profiles. The N2 O concentrations in both gas and liquid samples were analyzed using a gas chromatograph (GC) (Agilent 7820, USA) [12]. Acetic acid (HAc) and propionic acid (Pro) were measured according to the method described by Wang et al. [12]. PHB, PHV and PH2MV were measured according to the method described by Oehmen et al. [25]. The total PHA in the samples was calculated as the sum of the measured PHB, PHV and PH2MV. Dehydrogenase activity was determined according to the method described by Goel et al. [26], with minor modification. 3. Results and discussion 3.1. Cyclic studies in An/A/O SBR with acetate as a sole carbon source Cyclic studies were conducted in the An/A/O SBR to establish steady state operation and determine the reactor phenotypes. Fig. 2 shows a typical example of the concentration profiles measured during an SBR cycle after stable N and P conversions were achieved. The An/A/O SBR displayed a typical DPAOs phenotype.
Fig. 2. Variations in N, P, N2 O, HAc, glycogen and PHA in one cycle in the A/An/O SBR.
During the anaerobic phase, the acetate was typically completely consumed within 60 min, and this was accompanied by the formation of PHA, release of phosphorus and consumption of glycogen. In the subsequent anoxic period, once KNO3 solution was added to the reactor, simultaneous phosphorus uptake and denitrification occurred rapidly, concurrently with PHA degradation and glycogen replenishment. NO3 − –N was depleted after 60 min of anoxic operation. The NO2 − –N increased to 20 mg/L at 120 min, after which it was gradually reduced. Overall, approximately 2.60 mg N/L of N2 O was produced during the anoxic phase, giving a ratio of N2 O–N production to TN removal of 7.77% (Table 1). Regarding the post oxic phase, a small amount of NH4 + –N was oxidized to NO3 − –N or Table 1 Comparison of N and P removal and N2 O production in one cycle in the A/An/O SBR. Item
Values
MLSS (mg/L) MLVSS (mg/L) PO4 3− –P removal efficiency at the end of anoxic period (%)a PO4 3− –P removal efficiency at the end of oxic period (%)b TN removal efficiency (%)c Anoxic N2 O production (mg N/L) Oxic N2 O production (mg N/L) Ratio of anoxic N2 O–N production to TN removal (%)
3483 2002 82% 89% 77% 2.60 0.36 7.77%
a Anoxic end PO4 3− –P removal efficiency (%) = (P0 − P1 )/P0 ; P0 : PO4 3− –P concentration at the beginning of the anaerobic phase; P1 : PO4 3− –P concentration at the end of the anoxic phase. b Oxic end PO4 3− –P removal efficiency (%) = (P0 − P2 )/P0 ; P0 : P concentration at the beginning of the anaerobic phase; P2 : PO4 3− –P concentration at the end of the aerobic stage. c TN removal efficiency (%) = (TN1 − TN2 )/TN1 ; TN1 : TN concentration at the beginning of the anoxic phase; TN2 : TN concentration at the end of the anoxic phase.
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Table 2 Anaerobic and anoxic transformations of PHA and glycogen in one cycle in the A/An/O SBR. Item
Values
PHB synthesis (mmol-C/g-MLVSS) PHV synthesis (mmol-C/g-MLVSS) PHA synthesis (mmol-C/g-MLVSS) PHA synthesis/Hac uptake (mmol-C/mmol-C) Phosphorus release/Hac uptake (mmol-P/mmol-C) Glycogen degradation (mmol-C/g-MLVSS) PHB degradation (mmol-C/g-MLVSS) PHV degradation (mmol-C/g-MLVSS) PHA degradation (mmol-C/g-MLVSS) Glycogen synthesis (mmol-C/g-MLVSS) PHB synthesis/PHA synthesis (mmol-C/mmol-C) PHV synthesis/PHA synthesis (mmol-C/mmol-C) Hac uptake rate (mmol-C/g-MLVSS·h)
2.43 0.98 3.41 0.86 0.07 2.15 2.08 0.67 2.77 1.70 71% 29% 3.94
NO2 − –N (Fig. 2a), and this was accompanied by PHA degradation, glycogen replenishment and phosphorus assimilation (Fig. 2b). Moreover, approximately 0.36 mg N2 O–N/L was produced during this stage (Table 1), possibly via the denitrification pathway [12,27]. Kim et al. [27] reported that N2 O was generated in the aerobic nitrification stage when NO2 − –N and NH4 + –N existed simultaneously, and they deemed that N2 O was produced via the denitrification pathway by ammonia-oxidizing bacteria (AOB). It should be noted that the ratio of P release to Hac uptake for the An/A/O SBR was only 0.07 mmol-P/mmol-C (Table 2), which was significantly smaller than the ratios of 0.2–0.5 mmol-P/mmolC reported in previous studies [12,14,15,28]. Moreover, a high amount of PHV (29% of PHA) was produced during the anaerobic phase (Table 2). Taken together, these findings appear to indicate that glycogen accumulating organisms (GAOs) were present and active in the system and contributed to carbon source storage. GAOs can take up VFAs to store PHA under anaerobic conditions
by using glycogen as the intracellular energy pool instead of poly-P and not accumulating polyphosphate under aerobic/anoxic conditions, which results in competition with PAOs. Coincidently, GAOs tend to produce more PHV than PAOs via the propionate–succinate pathway [29]. Based on the long-term operation of the An/A/O SBR, the average effluent PO4 3− –P was high (3.49 mg/L) and the average PO4 3− –P removal efficiency was only 77% (data not shown), which confirmed that GAOs had survived in the SBR. The enrichment of GAOs in the studied SBR was primarily because of the high temperature (25–30 ◦ C), as temperatures over 20 ◦ C favor the growth of GAOs [30].
3.2. Shock effects of different carbon sources on denitrifying phosphorus removal and N2 O production in three batch reactors Batch tests were conducted with acetate-enriched biomass in the An/A/O SBR using acetate, mixed acetate and propionate, and propionate alone as the carbon sources.
3.2.1. Comparison of VFAs uptake in R-A, R-M and R-P As shown in Fig. 3f, the VFAs were completely taken up in R-A and R-M, while a fraction of propionic acid still remained in R-P at the end of the anaerobic stage. Additionally, the Hac uptake rate was higher in R-A than in R-M, and the propionic acid uptake rate increased by 2.95 times in R-M when compared to R-P (Table 3). Moreover, the propionic acid uptake rate was 1.26 times that of the Hac uptake rate in the R-M system. These findings suggest that the presence of Hac stimulated propionic acid uptake, while the presence of propionic acid inhibited Hac uptake. Similar results were also obtained by Li et al. [16], who examined the long-term effects of carbon source in an anaerobic–aerobic (low DO) phosphorus and nitrogen removal system and found that both propionic acid and
Fig. 3. Variations in N, P, N2 O and VFAs in R-A, R-M and R-P.
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Table 3 Anaerobic and anoxic transformations of PHA and glycogen in batch experiments. Item
Hac (R-A)
PHA synthesis/VFA uptake (mmol-C/mmol-C) Anaerobic end PHB (mmol-C/g-MLVSS) Anaerobic end PHV (mmol-C/g-MLVSS) Anaerobic end PHA (mg COD/L) PHB synthesis/PHA synthesis (mmol-C/mmol-C) PHV synthesis/PHA synthesis (mmol-C/mmol-C) Hac uptake rate (mmol-C/g-MLVSS h) Pro uptake rate (mmol-C/g-MLVSS h)
1.01 2.90 1.18 311 74.21% 25.66% 6.30 –
± ± ± ± ± ± ±
Hac uptake rates decreased as the ratio of propionic to acetic acid increased. Notably, the propionic acid uptake rates in R-M and R-P were 72% and 24% of the Hac uptake rate in R-A, respectively (Table 3). These findings indicate that biomass acclimatized with acetate may not be able to anaerobically take up propionic acid as quickly as Hac after only one cycle. Similarly, Oehmen et al. [15] found that the propionic acid uptake rate declined after a switch in carbon source from Hac to propionic acid, and that it decreased more as the fraction of GAOs increased. 3.2.2. Effects of carbon sources on the transformation of PHA and glycogen Fig. 4 shows the transformations of PHA and glycogen in R-A, R-M and R-P. When the carbon source was switched from acetate (R-A) to acetate/propionate (R-M) and then to propionate (R-P), the PHA synthesis decreased from 3.59 to 2.01 and then to 1.11 mmolC/g-MLVSS (Fig. 5a). Moreover, a decrease and an increase occurred
Glycogen (mol-C/g-MLVSS)
9
anoxic
anae robic
8
a)
7 6 Hac Hac+Pro Pro
5 4
PHA, PH2MV (mmol-C/g-MLVSS)
PHB, PHV (mmol-C/g-MLVSS)
3 5
b)
4
PHB (Hac) PHV (Hac) PHB(Hac+Pro) PHV (Hac+P ro) PHB (Pro) PHV (Pro)
3 2 1 0 5
c)
4
PHA (Hac) PH2MV (Hac) PHA (Hac+P ro) PH2MV (Hac+Pro) PHA (Pro) PH2MV (Pro)
3 2 1 0 0
30
60
90
12 0
15 0
18 0
210
24 0
270
time (min) Fig. 4. Variations of glycogen and PHA in R-A, R-M and R-P.
30 0
Hac + Pro (R-M) 0.03 0.08 0.16 19 1.71% 1.85% 0.24
0.61 0.46 2.06 203 14.59% 85.31% 3.60 4.55
± ± ± ± ± ± ± ±
0.01 0.15 0.23 30 0.59% 0.85% 0.01 0.23
Pro (R-P) 0.44 0.30 1.39 140 9.92% 89.61% – 1.54
± ± ± ± ± ±
0.06 0.07 0.20 22 2.02% 1.54%
± 0.33
in the fractions of formed PHB and PHV, respectively, as the percentage of added propionate increased (Table 3). These findings corresponded with those of Li et al. [16], who investigated an anaerobic/aerobic (low DO) process. Moreover, according to Lemos et al. [31], the primary fractions of PHA produced by PAOs were PHV and PH2MV when the carbon source was propionate. However, in the R-M and R-P systems, negligible amounts of PH2MV were anaerobically synthesized. Pijuan et al. [32] reported a similar finding, in which PH2MV was not synthesized when the carbon source was switched from acetate to propionate. However, further research is needed to clarify this issue. The ratios of PHA synthesis to VFA uptake (mmol-C/mmol-C) were lower in R-M and R-P (Table 3). This finding is in accordance with the results of several other studies [16,32] that revealed that propionate addition led to lower PHA synthesis/VFA uptake. It is also possible that some forms of PHA were not quantified, e.g. poly3-hydroxy-2-methylbutyrate (PH2MB). Lemos et al. [31] showed that synthesized PH2MB comprised 8.6% of the PHA formed when propionic acid was used as a carbon source by using 13 C-labeling and in vivo nuclear magnetic resonance. Moreover, it should be noted that about 0.6 mmol-C/g-MLVSS of PHA was degraded during the latter part of the anaerobic phase in R-M (Fig. 4b and c), leading to a lower net PHA synthesis/VFA uptake in R-M. Similar findings have been reported in our previous studies [12]. The amount of anaerobic glycogen degradation decreased when the carbon source was changed from acetate (R-A) to mixed acetate and propionate (R-M) and propionate (R-P), which corresponded well with the trend of the amount of PHA synthesis in each reactor (Fig. 5a and c). These findings were supported by the observations of Chen et al. [14] and Li et al. [16], who found that the increase in the influent ratio of propionic to acetic acid resulted in a decrease in glycogen degradation. During the anoxic phase, the amounts of PHA degradation in R-A, R-M and R-P were 3.94, 2.42 and 1.45 mmol-C/g-MLVSS, respectively (Fig. 5b), suggesting that increasing propionic acid content results in decreased PHA oxidation. In addition, the glycogen transformation showed a trend similar to the PHA degradation (Fig. 5b and c). 3.2.3. Comparison of N and P removal performance in R-A, R-M and R-P Of R-A, R-M, and R-P, which utilized acetate, acetate/propionate and propionate as the carbon source, respectively, the highest and lowest P release were observed in R-A and R-M, respectively (Figs. 3e and 5d). Moreover, anaerobic P uptake occurred in R-A (about 5.0 mg/L) and R-M (about 10.6 mg/L) during the latter part of the anaerobic phase, leading directly to a reduced net P release [12]. Therefore, the actual ability of P release was highest in R-A and lowest in R-P when acetate and propionate were used as the sole carbon source. This coincides well with other studies that have suggested that an increase in the propionic/acetic ratio led to less P release [16]. After the three reactors were fed with KNO3 solution, denitrification and P uptake occurred simultaneously. The highest P
1004
Y. Wang et al. / Chemical Engineering Journal 172 (2011) 999–1007
4.0
5.0
a
b
PHB synthesis PHV synthesis PHA synthesis
3.5
4.0 3.5 mmol-C/g-MLVSS
mmol-C/g-MLVSS
3.0 2.5 2.0 1.5 1.0
3.0 2.5 2.0 1.5 1.0
0.5
0.5 0.0
0.0 Hac (R-A)
3.0
Hac+Pro (R-M)
Hac (R-A)
Pro (R-P)
60
c
glycogen degradation glycogen syhtnesis
50
2.0
1.5
1.0
0.5
d
Hac+Pro (R-M)
phosphorus release phosphorus uptake TN removal
Pro (R-P)
100 phosphorus removal efficency TN removal efficency
90 80 70
40
60
30
50 40
20
30
removal efficency (%)
nitrogen or phosphorus (mg/L)
2.5 mmol-C/g-MLVSS
PHB degradation PHV degradation PHA degradation
4.5
20
10
10
0
0.0 Hac (R-A)
Hac+Pro (R-M)
Pro (R-P)
0
R-A (Hac)
R-M (Hac+Pro)
R-P (Pro)
Fig. 5. The effects of carbon sources on N and P removal and the transformations of glycogen and PHA in R-A, R-M and R-P.
removal was obtained in R-A, while the lowest was observed in R-P (Fig. 5d). This is in good agreement with the observation that a higher propionic acid content led to detrimental P removal in the short term in an anaerobic–oxic SBR [14]. Moreover, the added NO3 − –N was rapidly reduced in R-A, R-M and R-P, and this reduction was accompanied by the accumulation and reduction of NO2 − –N (Fig. 3b and c), which resulted in corresponding TN removal efficiencies of 80%, 52% and 38%, respectively. These findings indicate that, as the amount of propionate added increased, the denitrifying phosphorus removal decreased, based on only one cycle. 3.2.4. Effects of carbon sources on N2 O production in R-A, R-M and R-P Fig. 3d shows the variations in off-gas, dissolved and total N2 O concentrations in R-A, R-M and R-P. During the anaerobic phase, no N2 O production was detected in any of the reactors (Fig. 3d). However, after the KNO3 solution was pulse added at the end of the anaerobic period, a remarkable increase in the N2 O concentration occurred in all reactors. Thereafter, the N2 O production rate decreased gradually, and this was followed by a decrease in the N2 O concentration (Fig. 3d). Consequently, the ratios of N2 O–N production to the TN removal were 13.93 ± 0.05%, 37.95 ± 5.62% and 24.71 ± 4.59%, respectively, in R-A, R-M and R-P (Table 4). Obviously, when the carbon source was switched from acetate (R-A) to acetate/propionate (R-M) and propionate (R-P), the ratio of N2 O–N production to TN removal increased by 1.72 and 0.77 times, respectively. These findings suggest that changing the carbon source could lead to more N2 O production in one cycle during the denitrify-
ing phosphorus removal process. Moreover, the dynamic changes in carbon source types could produce a shock loading effect on the biomass in wastewater treatment bioreactors, which tends to result in poor nutrient removal efficiencies and a large amount of N2 O production. These scenarios also occur in practical wastewater treatment processes. 3.3. Factors leading to increased N2 O production after a change in carbon sources 3.3.1. Effect of activities of denitrifying enzymes The carbon source may affect the activities of denitrifying enzymes, leading to a different amount of N2 O production [13,17,19,21]. As shown in Table 4, when the carbon source was switched from acetate to acetate/propionate or propionate, the rates of NO3 − –N and NO2 − –N reduction both decreased, suggesting that the activities of nitrate and nitrite reductase were decreased. Furthermore, the carbon source shock can produce different effects on nitrate and nitrite reductase, which led to different ratios of the NO3 − –N reduction rate to the NO2 − –N reduction rate in R-A, RM and R-P (Table 4). Moreover, the NO3 − –N reduction rate was much higher than the NO2 − –N reduction rate (Table 4), and this imbalance caused a serious accumulation of NO2 − –N throughout the anoxic stage in the three reactors. This NO2 − –N accumulation increased as the ratio of the NO3 − –N reduction rate to the NO2 − –N reduction rate increased (Table 4 and Fig. 3c). Nitrite and FNA are known to inhibit the activity of nitrous oxide reductase (Nos), which causes the accumulation of N2 O in denitrification processes [10,11]. As a result, the greatest and lowest ratios of N2 O–N
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Table 4 Comparison of N and P removal and N2 O production in batch experiments. Item
Hac (R-A)
Hac + Pro (R-M)
Pro (R-P)
MLSS (mg/L) MLVSS (mg/L) NO3 − –N reduction rate (mg-N/g-MLVSS h) NO2 − –N reduction rate (mg-N/g-MLVSS h) NO3 − –N reduction rate/NO2 − –N reduction rate N2 O–N reduction rate (mg-N/g-MLVSS h) Dehydrogenase activity (mg TFa /g MLVSS h) PO4 3− –P uptake rate (mg-P/g-MLVSS h) Anoxic N2 O production (mg N/L) Ratio of anoxic N2 O–N production to TN removal (%)
3578 ± 49 2075 ± 68 16.60 ± 0.67 7.38 ± 0.95 2.25 5.73 ± 1.10 26.04 ± 2.53 19.82 ± 1.52 4.87 ± 0.13 13.93 ± 0.05
3619 ± 42 2117 ± 28 16.00 ± 0.76 5.56 ± 0.90 2.88 2.16 ± 0.32 21.08 ± 0.90 12.53 ± 0.40 7.57 ± 1.23 37.95 ± 5.62
3741 ± 62 2180 ± 79 11.71 ± 0.89 4.20 ± 0.37 2.79 1.68 ± 0.26 19.20 ± 1.53 8.18 ± 0.23 4.17 ± 1.27 24.71 ± 4.59
a
TF, triphenylformazan.
production to TN removal were observed in R-M and R-A, respectively, as the highest and lowest ratios of the NO3 − –N reduction rate to the NO2 − –N reduction rate occurred in R-M (2.88) and R-A (2.25), respectively (Table 4). Notably, the N2 O reduction rate in R-M and R-P was much lower than that in R-A (Table 4), suggesting that the activities of Nos were lower in R-M and R-P than in R-A. This further explained why the amount of N2 O produced by R-M and R-P was higher than that produced by R-A.
3.3.2. Effects of the amount and composition of anaerobically synthesized PHA Carbon source concentration is an important factor influencing N2 O production under anoxic conditions [33]. In our study, the carbon source for denitrification was from anaerobically synthesized PHA. Therefore, the amount of anaerobically synthesized PHA played an important role in N2 O production [12]. As shown in Table 3 and Fig. 5a, the levels of anaerobically synthesized PHA in R-A, R-M and R-P were 3.59, 2.01 and 1.11 mmol-C/g-MLVSS, corresponding to 311, 203 and 140 mg COD/L of PHA at the end of anaerobic stage, respectively. Obviously, the lower COD availability in R-M and R-P during the initial anoxic phase led to more N2 O production when compared to R-A. Additionally, even though more COD was available in R-M than in R-P, more N2 O was produced in R-M than in R-P. A possible explanation for this is that the propionate that was not consumed during the anaerobic period was residual to the anoxic phase in R-P, during which it was exhausted by ordinary denitrifying bacteria, which improved the denitrification rate and reduced the N2 O emissions [12]. Nevertheless, more NO2 − –N was accumulated in R-M than in R-P (Fig. 3c), which also attributed to the higher N2 O production in R-M than in R-P. Carbon source types have been reported to have impacts on N2 O emissions during biological denitrification [19–21]. In the present study, an internal carbon source (PHA) was used for biological denitrification, and the different composition of PHA (mainly PHB and PHV in the present study) may also be an important factor influencing N2 O production. As shown in Fig. 5a, the highest and lowest amount of PHV was formed in R-M and R-A, respectively, which corresponded to the greatest and lowest ratio of N2 O–N production to TN removal. These findings suggest that more PHV leads to higher N2 O production. As shown in Fig. 6, the metabolic pathway involved conversions of PHV in DPAOs that differed from those of PHB, as reported by Martin et al. [34]. Consequently, the organisms (PAOs, DPAOs or GAOs) may not be able to anoxically metabolize PHV as quickly as PHB in a transition period of only one cycle [15], which subsequently reduced the activities of the PAOs/DPAOs or GAOs. Moreover, the dehydrogenase activity, which reflects the microbial activity [26], declined after a switch in carbon source from acetate to acetate/propionate or propionate (Table 4). Similarly, the PO4 3− –P uptake rate and the NO3 − –N and NO2 − –N reduction rates
all decreased when the influent carbon source was switched from acetate (R-A) to mixed substrates (R-M) and then to propionate (RP) (Table 4). Taking into account all of these factors, it is clear that the use of PHV as a carbon source could result in a decrease in PAOs or GAOs activity after a transition in carbon sources in the short term, thereby triggering increased N2 O production. However, it is difficult to determine if the lack of acclimation of the biomass to these substrates was the reason for the high N2 O emissions, or if they were the normal N2 O emissions for these substrates. Accordingly, the long-term effects of PHV on N2 O production require further study. 3.3.3. Effects of oxidative stress on N2 O production in R-A, R-M and R-P The induction of oxidative stress in biological systems has been reported under unfavorable environmental or stress conditions [35,36]. Oxidative stress may cause detrimental effects in many cell types, and an oxidative stress response in bacteria has been implicated as the cause of process failure in wastewater treatment facilities. Moreover, oxidative stress could induce activation of the antioxidant response element (ARE), leading to the expression of cytoprotective enzymes, such as superoxide dismutase (SOD), which was detected by Martin et al. [34] in EBPR sludge. Superoxide dismutase catalyzes the reduction of superoxide anions into H2 O2 [37].
Fig. 6. Schematic diagram of the anoxic metabolic pathway for conversion of PHB and PHV in DPAOs.
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Although no studies have confirmed that oxidative stress occurs when the biomass being treated is subjected to a sudden switch in carbon source, we speculate that oxidative stress conditions may occur in our studied R-M and R-P systems. As shown in Fig. 3b, when the carbon source was switched from acetate to acetate/propionate or propionate, the NO3 − –N concentrations increased during the latter part of the anoxic period in R-M and R-P, indicating that nitrification reactions occurred. Conversely, this phenomenon was not observed in R-A. Therefore, we assume that trace of O2 was present in R-M and R-P. In the batch tests, the reactors were properly sealed off from the outside atmosphere and the DO level during the entire anoxic period was lower than 0.1 mg/L. However, oxidative stress associated with reduction of H2 O2 can produce O2 [36,37]. Sabumon P C also observed the anoxic nitrification, and suggested that trace of O2 was produced from the reduction of H2 O2 [36]. These findings showed the occurrence of oxidative stress in R-M and R-P. Furthermore, oxidative stress has adverse effects on DPAOs, DGAOs and enzymes [35], leading directly to increased N2 O production in R-M and R-P when compared to R-A. 4. Conclusions Carbon source shock significantly influenced N2 O production during the denitrification phosphorus removal process. The N2 O–N production to TN removal ratios were 1.72 or 0.77 times higher, respectively, when biomass acclimatized with acetate was subjected to acetate/propionate or propionate shock than when acetate alone was used. The shock change of carbon sources significantly influenced the activities of denitrifying enzymes and the amount and composition of anaerobically synthesized PHA, and also caused oxidative stress. This directly stimulated N2 O production. In addition, the TN removal efficiency decreased from 80% to 52% or 38%, respectively, and the PO4 3− –P removal efficiency declined from 75% to 63% or 47%, respectively, after the carbon source was switched from acetate to acetate/propionate or propionate. Therefore, acetate/propionate or propionate was unfavorable for conducting denitrifying phosphorus removal in only one cycle.
Acknowledgments This study was supported by the National Natural Science Foundation of China (NSFC) (no. 51078283), the Program for New Century Excellent Talent in University (no. NCET-08-0404), the Shanghai Science and Technology Development Funds (no. 09QA1406100) and the Fundamental Research Funds for the Central University (Tongji University).
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