Environ Sci Pollut Res DOI 10.1007/s11356-014-2837-4

RESEARCH ARTICLE

The influence of SBR parameters on the sludge toxicity of synthetic wastewater containing bisphenol A Xiurong Chen & Jianguo Zhao & Jun Zhao & Na Yang & Fei Zhang & Zijian Jiang

Received: 19 October 2013 / Accepted: 25 March 2014 # Springer-Verlag Berlin Heidelberg 2014

Abstract Synthetic wastewater with bisphenol A (BPA) concentrations of 7.5, 20, and 40 mg/L was treated with activated sludge sequential batch reactors (SBRs). The sludge acute toxicity indicated by the inhibitory ratio to luminous bacteria T3 was evaluated. The influent COD was controlled at approximately 300 mg/L, and aerobic conditions were maintained in the SBR. It was found that the process of BPA biodegradation, as opposed to BPA adsorption, contributed to the formation of sludge toxicity; there was a positive relationship between sludge toxicity and influent BPA concentration, and the toxicity centralized in intracellular regions and the intersection of extracellular polymeric substances (EPS) in sludge flocs. Since the BPA biodegradation process dedicated to sludge toxicity, the influence of key operational parameters such as sludge retention time (SRT) and hydraulic retention time (HRT) on sludge toxicity were investigated. It was founded that sludge toxicity decreased significantly when SRT and HRT were shortened from 20 to 10 days and 12 to 8 h, respectively. The results of Pearson correlation analysis indicated that the Shannon index H of the bacterial community correlated significantly to sludge toxicity. The results from both similarity analysis and UPGMA indicated that influent quality characteristic contributes much more to bacterial community than operation parameters, and then leads to difference between blank and control sludge toxicity. Keywords Wastewater . Bisphenol A (BPA) . Sequential batch reactor (SBR) . Sludge organic toxicity . Operation parameters Responsible editor: Ester Heath X. Chen (*) : J. Zhao : J. Zhao : N. Yang : F. Zhang : Z. Jiang School of Resource and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, People’s Republic of China e-mail: [email protected]

Introduction Land application is an important means of excess sludge disposal due to the high content of plant nutrient components (Singh and Agrawal 2008; Senesi et al. 2007; Andrés et al. 2011). However, sludge toxicity inhibits sludge reuse in soil. With regard to organic and inorganic sludge toxicity, the latter due mainly to heavy metals is well known, but organic toxicity has not been widely recognized. Nevertheless, it is not sufficient to assess ecological hazard only based on heavy metal toxicity for sludge land disposal (Alvarenga et al. 2007; Abad et al. 2005). Therefore, it is necessary to take organic toxicity into account when sludge reuse in soils is being evaluated. Typically, the relatively high content of toxic organics in sludge leads to organic toxicity. There are as many as 516 types organic compounds in sludge (Harrison et al. 2006), and most of the persistent organic pollutants (POPs) are introduced to the food chain connecting humans and ecosystems via sludge application to land (Clarke and Smith 2011). In excess sludge in Britain, Spain, Australia, and China, polychlorinated-p-two benzene dioxins/furans (PCDD/Fs), phthalate esters (PAEs), polybrominated diphenyl ethers (PBDEs), polybrominated biphenyls (PBBs), polycyclic aromatic hydrocarbons (PAHs), benzo[a]pyrene (B[a]P), and polychlorinated biphenyls (PCBs) have been detected (Bright and Healey 2003; Gomez-Rico et al. 2008; Clarke et al. 2008a; Clarke et al. 2008b; González et al. 2010; Cantarero et al. 2012; Cai et al. 2007; Zeng et al. 2012; Hua et al. 2008; Shen et al. 2009; Dai et al. 2007; Yang et al. 2011). In addition to the final biodegradation products of pollutants, the biodegradation process itself contributes to the formation of sludge toxicity. Some researchers have demonstrated that stress effects from toxic pollutants stimulate the microorganisms to produce toxic soluble microbial products (SMP) (Yu et al. 2006; Xu J. 2010) and stress protein GroEL (Bottm and

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Love 2001). When the Daphnia magna Straus is taken as a toxicity indicator, it is seen that wastewater containing toxic organics led to the formation of sludge toxicity indirectly (Yu et al. 2004). To investigate toxicity formation and reduction in sludge, the relationship between synthetic wastewater containing toxic bisphenol A (BPA) treatment and sludge toxicity was studied in sequential batch reactors (SBRs). BPA was chosen as the target pollutant based on its properties as follows: (1) BPA is a widely used industrial material (Deng and Wu 2004; Craig et al. 2003) that is ubiquitous in industrial wastewater. Therefore, BPA treatment is representative of toxic wastewater treatment processes. (2) BPA is biodegradable (IKE 1999) and toxic (Can et al. 2005). More importantly, BPA has low volatility and low bioaccumulativity (Wang and Zhao 2007). Thus, there is no interference for analyzing BPA removal routes except between adsorption or biodegradation and sludge toxicity formation. Before the present study, literature concerning BPA biological treatment has been dominated by BPA removal from effluent, dominant bacteria cultivation, and BPA degradation pathways (Chen et al. 2008). The process of toxicity formation and sludge disposal for treating wastewater containing BPA has not been documented; this is the key point in this research. Due to the complicated constituents of real wastewater, it will disturb assessing the mechanism of toxicity formation and spatial distribution. So the influents were prepared synthetically. Otherwise, many other influent toxic constituents in real wastewater such as heavy metals, organics with higher toxicity than BPA, would lead to more significant toxicity of sludge. And then, when BPA is added into the real wastewater, the sludge toxicity would be mainly contributed by other toxic substances in raw wastewater. In order to investigate toxicity spatial distribution in sludge, the toxicity of outer section EPS in sludge flocs was considered. The stable characteristic of EPS depends on the steady influent quality, which is benefit to examining toxicity distribution in sludge flocs. In order to evaluate the contribution of BPA biodegradation to sludge toxicity and toxicity spatial distribution, the synthetic wastewater groups with and without BPA were contrasted for toxicity variation in sludge.

control group. To analyze sludge toxicity formation without interference from various pollutants in real wastewater, the wastewater was prepared synthetically. The blank SBR was fed only with peptone as organic substrate, whereas the control SBR was fed with both peptone and BPA. The influent COD of both blank and control SBRs was approximately 300 ±20.5 mg/L. The ratio of original C: N: P was 100:5:1 to meet the nutrient requirement of aerobic activated sludge. Since the real concentration of BPA in the influent of some wastewater plants is up to 825 μg/L (Mohapatra et al. 2010), that in landfill leachate is about 17.2 mg/L (Andrew Crain et al. 2007). In some, researchers also take the target BPA concentrations from 0.1 to 15.8 mg/L (Chen Jianhua et al 2008). Based on the information from relevant literature, the BPA original concentration in this research was taken as 7.5, 20, and 40 mg/L to investigate the influence of BPA load on sludge toxicity. For the original BPA concentration variation from 7.5±1.0 mg/L through 20±1.0 mg/L to 40±1.0 mg/L, there were three phases of study, each lasting 30 days. The initial aerobic activated sludge was from the Chang Qiao wastewater treatment plant in Shanghai, China. The 12-h period of the SBR process consisted of feeding (10 min), aeration (8 h), stirring (3 h), settling (1 h), and decanting (10 min). The dissolved oxygen (DO) was maintained at 2.0–3.0 mg/L throughout the 12-h biological process. The microbial suspended solid (MLSS) concentration was kept at approximately 3,000±100.0 mg/L. To investigate the influence of operation parameters on sludge toxicity, sludge toxicity under conditions of long and short HRT and SRT were compared. For the long times, HRT and SRT were 12 h and 20 days, respectively; for the short times, HRT and SRT were 8 h and 10 days, respectively. Analytical methods COD analysis After SBR reactor was settled, the supernatant was centrifuged at 4,000 r/min for 10 min to analyze COD, and the COD was measured in parallel according to APHA, 2002. The LOQ of COD determination was 0.01 mg/L. One hundred fifty milligrams per liter COD of phthalic acid standard solution was measured three times in laboratory and the repeatability was 147.7 mg/L±5.1 %. Plus-minus represented relative standard deviation (RSD). BPA analysis

Materials and methods Reactor design and synthetical wastewater The contrast conditions were tested in duplicate in two 10 L bench-scale SBRs, which were defined as the blank group and

BPA mainly included two parts, which were both in aqueous and sludge phase. After the sample was pretreated, aqueous phase and sludge phase BPA were measured in parallel by high-performance liquid chromatography (HPLC, LC10ATVP, Kyoto, Japan) using a reversed-phase C-18 column

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(250 nm×4.6 nm, 5 μm) as stationary phase, and a mixture of methanol and H2O (77:23) as mobile phase. The flow rate was maintained at 1 mL/min, and a wavelength of 280 nm was used. The pretreatment methods were as follows: Fifty milliliter mixed liquor was obtained and centrifuged at 4,000 r/min for 10 min to separate the aqueous and sludge phases. The aqueous phase was extracted using 20 mL dichloromethane, and the sludge phase was ultrasonically extracted using 20 mL methanol and dichloromethane (1:1) twice. Then, the methanol or dichloromethane was evaporated using a rotary evaporator at 30 °C. Finally, the residues were redissolved with 2 mL methanol and measured by HPLC. The (limit of the detection) LOD of BPA detection with HPLC was 0.01 mg/L. The analytical conditions of BPA by HPLC are shown in Table 1. The LOD of BPA detection with HPLC was 0.01 mg/L. Ten milligrams per liter BPA standard solution was measured three times in laboratory, and the repeatability was 9.95 mg/L ±1.2 %. Plus-minus represented relative standard deviation (RSD). Test for recovery of BPA were made at three concentration levels (1, 5, and 10 mg/L) both in aqueous phase and sludge phase; values of extraction efficiency were found in the range of 83.2–90.6 % and 68.5–82.4 % with RSD (n=5) in the range of 1.4–3.2 % and 1.8–4.3 %. EPS extraction from the activated sludge sample EPS extraction procedure was based on the research conclusion of B Frølund in 1996 (Frølund et al. 1996). First, the activated sludge was mixed with cation exchange resin (CER). The CER was a Dowex 50×80, Na+ from, 20–50 mesh in the sodium form (Fluka 44445). The 50-ml sludge sample was centrifuged twice at 4,000g and once for 10 min, the supernatant was removed, and the sediment was rinsed with buffer solution. Dowex was scattered into the sediment sludge with 75 g Dowex per gram of volatile suspended solid (VSS). Then, the mixed liquor was stirred in a conical flask for 2 h at 4ºC. After stirring, the extracted EPS was separated by centrifugation at 12,000g and 4ºC for 10 min and stored for toxicity analysis [Gao et al. 2008].

Cells lysis The ultrasonic irradiation led to the breakup of sludge flocs, followed by the release of the cellular content into the supernatant (Cao et al. 2005). First, the sludge sediment was rinsed in a buffer solution after EPS was extracted. Second, the 50-ml sludge sample was filled into a plastic centrifuge tube and performed in a probe system that emitted 20 kHz (0.5 w/ml) power density of ultrasound through a tip for 10 min sonication. The ultrasonic probe was kept 3-cm deep in the sludge phase, and the centrifuge tube was fixed in an ice bath throughout the operation. The organic matter released from sludge flocs was centrifuged at 4,000g for 20 min. Then, the supernatant was obtained after centrifugation at 12,000g and 4ºC for 10 min [Hong et al. 2002; Gao et al. 2008]. Detection of sludge acute toxicity There are three kinds of toxicity evaluation methods such as microorganism toxicity, plant toxicity, and animal toxicity (Gallardo et al. 2012; Mantis et al. 2005; Ricco et al. 2004 ). The luminescent bacterium (Agnés et al. 2001) and daphnia magna (Xing et al. 2012) are usually for microorganism toxicity evaluation; the seed germination percent or rate (Hamdi et al. 2006) and growth characteristics of some kinds plant are taken as plant toxicity objects; Daphnia magna in soil are often chosen for animal toxicity analysis. In order to investigate the sludge toxicity extent and impaction to microbes, the luminescent bacterium, Photobacterium phosphoreum (T3 mutation) was taken as the study object in this research. The freeze-dried luminescent bacterium, P. phosphoreum (T3 mutation) from the Institute of Soil Science, Chinese Academy of Sciences, Nanjing, People’s Republic of China was reconstituted by 3 % NaCl (Li et al. 2012). Inhibition of T3 mutation was related directly to the acute toxicity of sludge samples. The methods for this test were based on the GB/T 15441-1995 standard analytical methods. The expression of relative inhibiting luminosity is as follows:

  luminosity of the sample tube relative inhibiting luminosity ð%Þ 1 100 luminosity of the CK tube

PCR-DGGE analysis The DNA in activated sludge was extracted using 3 s DNA isolation kits for environmental samples. The DNA fragments were displayed clearly by agarose gel electrophoresis and

were suitable for denaturing gradient gel electrophoresis (DGGE) analysis. The bacterial communities of sludge samples were analyzed by the DGGE process. In the PCR-DGGE analysis process, a universal primer pair, 341f (forward primer) and

Environ Sci Pollut Res Table 1 Analytical conditions of BPA by HPLC Target Mobile phase/ Flow rate Wavelength Temperature Injection (mL/min) (nm) (°C) volume (μL) object methanol: water BPA

77:23

1.00

280

30

20

907r (reverse primer), was used to amplify the bacterial V3 to V5 region of the 16s rDNA gene, with a GC-clamp attached to the forward primer. Each PCR mixture contained 1 μL extracted DNA as template, 2 μL forward and reverse primer, 5 μL Ex Taq buffer (4 mM Mg2+), 4 μL dNTP mixture (0.4 mM, respectively), and 0.25 μL TaKaRa Ex Taq. The volume of PCR mixture was calibrated to 50 μL with deionized water. The PCR amplification process was as follows: initial denaturation at 95ºC for 10 min, PCR mixture circulation 32 times at 95ºC for 30 s, annealing at 55ºC for 40 s, extension at 72ºC for 1 min, circulation for 30 times and storage at 72ºC for 10 min. The amplimer of 16s rDNA gene fragments obtained by PCR was verified by electrophoresis on 1.0 % agarose gels stained with ethidium bromide. DGGE analysis was performed with A DCode universal mutation detection system (Bio-Rad, USA) based on the Cremonesi technique (Cremonesi et al. 1997). The samples containing equal amounts of PCR products were loaded into linear porosity of 6 % denaturing gradient acrylamide that contained 30–50 % denaturing gradient gel (100 % denaturant is a mixture of 7 M urea and 40 % deionized formamide). Electrophoresis was performed at 200 V for 6 h in 1×TAE running buffer. Gels were stained with GelRed solution for 20 min and repeated three times. Then, the gel was photographed under UV transillumination in a Gel Doc 2000 system (Bio-Rad).

Results and discussion Sludge toxicity formation mechanism and property Variations of effluent COD and sludge toxicity with BPA initial content Various BPA initial concentrations (7.5, 20, and 40 mg/L) and the corresponding effluent COD and sludge toxicity of two SBRs are shown in Fig. 1. Each initial BPA concentration phase was 30 days, which exceeded the SRT (20 days). After the 15th day, the effluent COD was less than 55 mg/L. The effluent BPA concentration was below the detection limit of the HPLC (0.01 mg L−1), and MLSS concentrations of SBRs remained at approximately 3,000 mg L−1. Thus, the sludge characteristic seemed to achieve steady state. At each initial BPA concentration phase, the control sludge was acclimated

with the previous BPA concentration for more than one SRT, e.g., 30 days. So effluent COD of control SBR was relatively higher than that of blank SBR before 10 days for each phase. Because BPA is biodegradable with acclimated aerobic activated sludge (Chen et al. 2008), there was no evident difference between effluent COD of two SBRs after about 15 days as the sludge of control SBR was acclimated with the current BPA concentration. However, the sludge toxicity of the control group increased with BPA initial concentration. Contribution of adsorption or degradation of BPA to sludge toxicity Activated sludge is capable of adsorbing pollutants due to its high specific surface area (Leung et al. 2012). To investigate whether sludge toxicity was caused whether by biodegradation or adsorption, the toxicity, BPA adsorption, and biodegradation of three types of sludge were compared. The three types of sludge were blank SBR, control SBR, and inactivated blank SBR. The initial BPA concentration was 20 mg/L. A total of 300 ml of sludge with MLSS at 3,000 mg/L was averaged into three parts, and each portion was placed into 250-ml Erlenmeyer flasks with breathable gauze to maintain aerobic conditions. The amount of BPA in the sludge was defined as BPA removed via adsorption. The difference between the influent BPA and the BPA removed via adsorption was defined as BPA removed via biodegradation. The removal of BPA via various routes is illustrated in Table 2. Organic substrate removal by active sludge was proven to be caused by adsorption and biodegradation (Leung et al. 2012). So BPA removal from supernatant separately by active blank sludge and control sludge was due to adsorption and biodegradation, and the amount of BPA in inactivated sludge was due to adsorption. The inactivated blank sludge had more uniform structure than the other two kinds of sludge because the sludge flocs were broken in the process of inactivation. Wang et al reported that the specific surface area was increased when the solid size was smaller (Wang et al. 2007). So it was inferred that the inactivated blank sludge had high specific surface area than the other two kinds of sludge. Because the activity of the inactivated sludge was destroyed, BPA removal from supernatant was caused by inactivated sludge adsorption. The spatial distribution of sludge toxicity in sludge flocs is shown in Fig. 2. Because the properties of inactivated sludge exhibited uniformity, no distinct difference was observed between the toxicities of whole sludge flocs and outer section EPS. In contrast, for blank and control sludge, the inactivated sludge toxicity was almost 20 % less. BPA removal was due to adsorption for inactivated sludge, whereas the activated blank and control sludge achieved removal mostly via biodegradation. It is concluded that the adsorption of BPA on inactivated sludge contributed minimally to sludge toxicity.

Environ Sci Pollut Res Fig. 1 Variation of sludge toxicity and effluent COD with BPA initial concentration. Each COD concentration sets two parallel samples, and sludge toxicity sets three parallel samples. The plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

Since the stress protein GroEL was proven to be secreted by sludge under impaction (Bottm and Love 2001), then the characteristic of sludge was changed. Toxic SMP was also produced as toxic original pollutants existed (Wang and Zang 2010). The BPA had impaction on blank sludge, so the toxicity of blank sludge was caused by BPA. While the accumulation of toxic matter from BPA quick biodegradation resulted in control sludge toxicity. Although the factors for toxicity formation were different, the extent of inhibition to luminosity by two types of sludge toxicity was similar. The relatively low toxicity of outer EPS indicated that there existed little toxic substances. Based on the above results, it is concluded that the toxicity of sludge was mainly caused by BPA biodegradation rather than adsorption. In the process of BPA biodegradation, the products were 2hydroxyphenyl-2-ketone-1-ethanol, 2,3-bis(4hydroxyphenyl)-1,2-propanediol, 2-hydroxyphenyl-2-

ketone-1-ethanol, 2-hydroxyphenyl acetic, 2-hydroxy-3-styrene acrylic acid, 1-vinyl-4-methoxy benzene, and 1-vinyl-4methoxy benzene by sequence (J Spivack et al 1994; Soo-Min Lee et al 2005 ). Since the BPA metabolites are simple organic acid and organic alcohol, the toxicity of them are much lower than BPA. So some toxic matters secreted by bacterial community under BPA impaction were indirectly proven to be the key factor for sludge toxicity formation. Sludge toxicity reduction by changing the SBR parameters Because sludge toxicity was formed in the process of BPA biodegradation, the SBR parameters such as hydraulic retention time (HRT) and sludge retention time (SRT) relating directly to BPA biodegradation were regulated to reduce sludge toxicity. The initial BPA concentration of control SBR was 40 mg/L. With the exception of HRT and SRT,

Table 2 BPA removal via various sludge adsorption or biodegradation Sludge samples

BPA content/removal (mg/L)

Inactivated blank sludge BPA of supernatant BPA on inactivated sludge BPA removal by adsorption Blank sludge BPA of supernatant BPA on blank sludge Control sludge

BPA removal by biodegradation BPA of supernatant BPA on control sludge BPA removal by biodegradation

0 h 0.5 h

2h

4h

6h

8h

10 h

20 0 0 20 0

7.48±0.26 12.52±0.52 0 3.73±0.18 12.25±0.47

6.68±0.35 13.32±0.64 0 4.49±0.11 10.81±0.52

2.01±0.12 17.99±0.43 0 6.4±0.22 2.46±0.09

2.47±0.18 17.53±0.38 0 7.32±0.23 1.26±0.04

3.25±0.15 16.75±0.74 0 6.54±0.33 1.51±0.05

3.15±0.08 16.85±1.02 0 6.17±0.41 2.07±0.08

0 20 0 0

4.02 4.11±0.25 11.06±0.49 4.83

4.7 0.07±0.03 0.35±0.11 19.58

11.14 0.06±0.04 0.22±0.07 19.72

11.42 0.05±0.01 0.2±0.07 19.75

11.95 0.05±0.02 0.23±0.06 19.72

11.76 0.04±0.01 0.14±0.04 19.82

Each BPA concentration sets two parallel samples, and the plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

Environ Sci Pollut Res Fig. 2 Variation of toxicity and spatial distribution in various types of sludge flocs. Each sludge toxicity sets three parallel samples, and the plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

which were changed from 12 to 8 h and from 20 to 10 days, respectively, other parameters were kept constant.

Influence of parameters on liquid COD, BPA, and sludge toxicity in one SBR cycle

Influence of parameters on effluent COD, BPA, and sludge toxicity in the operation period

The variations of liquid COD and sludge toxicity in one SBR cycle are illustrated in Fig. 4 for the case of stable blank and control SBRs (e.g., at the 28th day after parameters were changed). The BPA in both liquid and sludge was below the detection limit of HPLC after 2 h for two operation modes. As shown in Fig. 4, COD fluctuations were negatively correlated with sludge toxicity, especially from 1 to 2 h. The fast-biodegradation bacteria selected by short SRT and HRT contributed to toxic intermittent substances accumulation in the beginning of the SBR cycle (1 to 2 h). So COD removal rate was slowed down and, in turn, toxicity increased significantly. After 2 h, with the fast consumption of intermittent products by dominant bacteria in sludge, sludge toxicity reduced sharply. The peak value time for sludge toxicity under original and new operation modes were at 8 and 2 h, respectively. It is inferred that the accumulation rate of toxic materials was different under various conditions. Because BPA was more recalcitrant to biodegradation and more toxic than peptone (IKE 1999; Can et al. 2005), the toxicity of control sludge was always above that of blank sludge.

There was little fluctuation for effluent COD after HRT and SRT were shortened. The effluent COD of both SBRs stayed in the range 50–60 mg/L. The effluent BPA concentrations of two SBRs were always below the detection limit of HPLC (Kyoto, Japan). The variation of sludge toxicity with different parameters is shown in Fig. 3. In contrast, the toxicities of blank sludge and control sludge decreased to approximately 9.0 and 15.0 %, respectively, when SRT and HRT were shortened. This phenomenon about sludge toxicity variation with SRT regulation was similar to research of Sponza et al. who reported that effluent toxicity was not inevitably decreased with SRT increment (Sponza and Gok 2011). Some researchers demonstrated that the slow growing bacteria were outcompeted by the fast-growing bacteria as SRT was shortened significantly (Majewsky et al. 2011; Winkler et al. 2012). So the fast-growing bacteria became the dominant bacterial community as SRT was shortened from 20 to 10 days. Furthermore, with HRT reduction from 12 to 8 h, the organic load of sludge increased by 50 %. As a result, the bacteria with short generation time and fast biodegradation rate were chosen as the dominant community. Thus, the toxic substances were consumed more quickly, and sludge toxicity of blank and control SBR was reduced greatly.

Fig. 3 Variation of sludge toxicity with different HRT and SRT. Each sludge toxicity sets three parallel samples, and the plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

Correlation between bacterial community and sludge toxicity with various parameters Digitized DGGE images were analyzed with Quantity One image analysis software (version 4.6, Bio-Rad, USA). Band matching was performed with 4.00 % position tolerance and

Environ Sci Pollut Res Fig. 4 Variation of effluent COD and sludge toxicity with different HRT and SRT in one SBR cycle. Each COD concentration sets two parallel samples, and sludge toxicity sets three parallel samples. The plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

1.00 % optimization. A cluster analysis of the four samples was performed based on the Pearson similarity correlation and the UPGAMA or WPGAMA mean dendrogram method, which was used to determine the similarity index of different DGGE profiles. Correlation between Shannon-Weaver diversity index and sludge toxicity Since the Shannon-Weaver diversity index (H) considers the number of species and the consistency of a given community (Shannon and Weaver 1949), H is a common-used index of biodiversity (Shannon and Weaver 1963). In this study, the index H calculated through DGGE band profile analyses was used to estimate bacterial diversity among various sludge samples. Each sludge sample was collected at the 28th day in different operation processes. The equation used to calculate H is: H=−∑(ni/N)ln(ni/N), where ni is the peak height of the band i, i is the number of bands in each DGGE gel profile, and N is the sum of peak heights in a given DGGE gel profile.

To investigate the bacterial community of four sludge samples, a DGGE analysis of the four sludge samples (blank sludge and control sludge under various operation modes) was conducted. Each sludge sample was collected at the 30th day in different operation processes, namely with SRT and HRT were 20 days, 12 h, and 10 days, 8 h, respectively. The information about the sludge samples and the results of the DGGE analysis are summarized in Table 3. In the period of 30 days after SRT and HRT were shortened, the average toxicity values of blank sludge and control sludge decreased from 38.58 to 13.42 % and from 48.89 to 32.58 %, respectively. The H of blank sludge increased from 3.59 to 3.88, and the H of control sludge reduced from 3.72 to 3.67. Under the original operation mode, the endogenous respiration of blank sludge was evident as COD in liquid was removed quickly to the minimum as 12.20 mg/L and the relevant organic substrate load was as low as 0.004 g COD/ g SS, which was indicated by the COD increment at the 8th hour in Fig. 4. Under the new operation mode, even though some low-growing bacteria were outcompeted by fast-

Table 3 Shannon-Weaver diversity index and sludge toxicity of four sludge samples Sample number

No. 1

Sludge samples

Blank sludge under original Blank sludge under new Control sludge under original Control sludge under new operation mode operation mode operation mode operation mode 38.58 %±3.57 % 13.42 %±2.14 48.89 %±5.62 % 32.58 %±6.23 %

Sludge toxicity

Shannon-Weaver diversity index (H) 3.59

No. 2

3.88

No. 3

3.72

No. 4

3.67

1. All the sludge samples were collected from SBRs at the 30th day of operation under original or new parameters. 2. The initial BPA concentration of control SBR was 40 mg/L. 3. Each sludge toxicity sets three parallel samples, and the plus-minus is standard deviation of parallel samples, which characterizes the dispersion of the determination results

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growing bacteria, the endogenous respiration was weakened because organic load increased 50 % when HRT was changed from 12 to 8 h, which was embodied as the continual reduction of COD. When the deaths from endogenous respiration were more than that from competition, H of blank sludge increased from 3.59 to 3.88. In the control SBR, the high strength BPA was more recalcitrant than peptone, so COD in liquid was removed continually under original operation mode due to negligible endogenous respiration. Therefore, bacteria did not die much from endogenous respiration under original operation mode. When SRT and HRT were reduced greatly, the fast-growing BPA and peptone biodegradation bacteria were chosen as the dominant community in control sludge, and then the toxicity decreased according to the fast consumption by the powerful dominant bacteria. As a result, with similarly weak endogenous respiration, some bacteria outcompeted with short SRT and HRT leading to biodiversity decreasing from 3.72 to 3.67. Correlation between similarity index and sludge toxicity The PCR-DGGE profile similarity between various sludge samples were calculated with Dice coefficient, which is shown in Table 4. As to the sludge samples with same influent property (namely blank sludge samples 1 and 2, control sludge samples 3 and 4), the DGGE profiles exhibit the relatively high similarity (71 % for control sludge samples and 63 % for blank sludge samples). However, either under original or new operation modes, the similarity between blank sludge and control sludge was as low as approximately 55 %. Cluster analysis of bacterial communities of various sludge samples was performed with the unweighted pair group method with arithmetic averages (UPGMA). The results are shown in Fig. 5. The four sludge samples are classified into two clusters on basis of UPGMA. The sludge samples in either cluster had the same influent quality. The similarity between two sludge clusters was as low as 0.57 and that between two blank sludge samples and two control sludge samples were, respectively, 0.63 and 0.71. The results from both similarity analysis and UPGMA indicate that influent quality property rather than operation parameters such as SRT and HRT wegre the key factor resulting to bacterial community alternation and sludge Table 4 Dice coefficients comparing the similarities of PCRDGGE fingerprints (%)

Fig. 5 Dendrogram of DGGE patterns of four sludge samples by UPGMA

toxicity difference. Therefore, optimizing SRT and HRT contributed to sludge toxicity reduction.

Conclusion Sludge toxicity formation mechanisms and toxicity reduction by adjusting parameters in SBR operation were investigated. It was found that the biodegradation process of BPA, as opposed to the adsorption of BPA, contributed to the formation of sludge toxicity. There was a positive correlation between sludge toxicity and original BPA concentration. The toxicity was predominantly distributed in the inner section of sludge flocs. When SRT and HRT were shortened from 20 to 10 days and 12 to 8 h, respectively, the fast-growing bacteria were the dominant community, and sludge toxicity decreased significantly. However, the results from both similarity analysis and UPGMA indicated that influent quality contributed more to bacterial community than operation parameters. In addition, the results of Pearson correlation analysis indicated that there existed correlation between sludge toxicity and the Shannon index H. When the bacterial community was influenced by the toxic target pollutant BPA, the sludge toxicity was accumulated during BPA biodegradation process. Therefore, the life activities of bacterial community impacted by influent BPA led to sludge toxicity indirectly, and in turn, SRT and HRT regulation caused sludge toxicity variation for the reason that the two operation parameters affected bacterial community characteristic. The sludge toxicity formation mechanisms and toxicity reduction by adjusting operation parameters are instructive for maintenance of real organic industry wastewater bio-treatment process.

Lane

No. 1

No. 2

No. 3

No. 4

No. 1 (blank sludge under original operation mode) No. 2 (blank sludge under new operation mode) No. 3 (control sludge under original operation mode) No. 4 (control sludge under new operation mode)

100.0 63.0 55.0 58.7

63.0 100.0 57.5 55.2

55.0 57.0 100.0 71.0

58.7 55.2 71.0 100.0

Environ Sci Pollut Res Acknowledgments This study was financially supported by the Natural Science Foundation of China (No. 51008124, 51378207), the Natural Science Foundation for Distinguished Young Scholars of China (No. 51125032), and the Shanghai Pujiang Program (No. 13PJD009) in China.

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