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Feb 19, 2010 - individual toxic effects of long-term addition of Cu(II) and. Ni(II) on the biochemical properties of aerobic granules in sequencing batch reactors ...
Appl Microbiol Biotechnol (2010) 86:1967–1975 DOI 10.1007/s00253-010-2467-9

ENVIRONMENTAL BIOTECHNOLOGY

Effects of long-term addition of Cu(II) and Ni(II) on the biochemical properties of aerobic granules in sequencing batch reactors Xin-Hua Wang & Li-Hong Gai & Xue-Fei Sun & Hui-Jun Xie & Ming-Ming Gao & Shu-Guang Wang

Received: 6 December 2009 / Revised: 22 January 2010 / Accepted: 22 January 2010 / Published online: 19 February 2010 # Springer-Verlag 2010

Abstract Copper (Cu(II)) and nickel (Ni(II)) are often encountered in wastewaters. This study investigated the individual toxic effects of long-term addition of Cu(II) and Ni(II) on the biochemical properties of aerobic granules in sequencing batch reactors (SBRs). The biochemical properties of aerobic granules were characterized by extracellular polymeric substances (EPS) content, dehydrogenase activity, microbial community biodiversity, and SBR performance. One SBR was used as a control system, while another two received respective concentration of Cu (II) and Ni(II) equal to 5 mg/L initially and increased to 15 mg/L on day 27. Results showed that the addition of Cu (II) drastically reduced the biomass concentration, bioactivity, and biodiversity of aerobic granules, and certainly deteriorated the treatment performance. The toxic effect of Ni(II) on the biodiversity of aerobic granules was milder and the aerobic granular system elevated the level of Ni(II) toxicity tolerance. Even at a concentration of 15 mg/L, Ni (II) still stimulated the biomass yield and bioactivity of aerobic granules to some extent. The elevated tolerance seemed to be owed to the concentration gradient developed within granules, increased biomass concentration, and promoted EPS production in aerobic granular systems.

X.-H. Wang : L.-H. Gai : X.-F. Sun : M.-M. Gao : S.-G. Wang (*) Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Jinan 250100, China e-mail: [email protected] H.-J. Xie Environment Research Institute, Shandong University, Jinan 250100, China

Keywords Aerobic granules . Cu(II) . Ni(II) . Toxic effects . Biochemical properties

Introduction Aerobic granulation technology is a novel environmental biotechnology and has been studied extensively (Beun et al. 2002; Liu et al. 2005; Wang et al. 2008; Juang et al. 2009). Aerobic granules have strong structure, excellent settleability, and high biomass retention. It has been used for the treatment of multiform wastewaters containing organics, nitrogen, and phosphorus (de Kreuk et al. 2005; Wang et al. 2007; Adav et al. 2009; Ni et al. 2009). Particularly, the aggregation of microorganisms into compact aerobic granules confers additional benefits such as protection against resistance to chemical toxicity (Jiang et al. 2004a; Yi et al. 2006; Adav et al. 2008; Zhang et al. 2008). Aerobic granules hereby have advantages over conventional activated sludge for treating wastewaters containing inhibitory and toxic substances, such as metal industrial wastewater. Heavy metals can be stimulatory, inhibitory, or even toxic in biochemical reactions depending on the metal concentration and speciation, the state of microbial growth, and the biomass concentration. It is reported that low concentrations of so-called “essential” metals (such as Fe, Zn, Cu, Ni, and Co) can stimulate biological systems (McCarthy 1964), while further increase in heavy metal concentration causes the inhibition and even failure of the biological treatment system (Yetis and Gokcay 1989). Different methods have been developed to quantify metal ion toxicity on activated sludge systems, such as inhibition of enzymatic activities, respiratory activities of the bacteria, kinetics of bacterial activities, and dynamics of microbial

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community evolution (Hu et al. 2002; Nicolau et al. 2005; Principi et al. 2006; You et al. 2009). However, to our knowledge, information about the effects of heavy metal ions on aerobic granular systems is not yet available. In addition, long-term effects of heavy metal ions addition need to be investigated because short-term batch assays may not be used directly to infer toxic effects in reactors with long-term metal exposure (Vandevivere et al. 1998; Hu et al. 2004). This study was thus conducted to examine the individual effects of long-term addition of Cu(II) and Ni(II) on the biochemical properties of aerobic granules in sequencing batch reactors (SBR). Cu(II) and Ni(II) were selected as model heavy metals in this study due to its common presence in metal industrial wastewater and both belong to the “essential” metals. Toxic effects of heavy metal ions on microbial community bioactivity and biodiversity of aerobic granules were analyzed by dehydrogenase activity (DA) and Biolog tests and multivariate analysis. It is expected that the information provided here would be useful to facilitate the application of aerobic granulation technology in metal industrial wastewater treatment.

settling velocity of 52 m/h. R1 was used as a control system in this study. When the systems were acclimatized to the feed, as indicated by a stable COD removal efficiency, CuCl2∙2H2O and NiCl2·5H2O solutions were added to synthetic wastewater to provide 5 mg/L Cu(II) and Ni(II) fed in R2 and R3 during the first 26 days (stage I), respectively. At stage II, concentrations of Cu(II) and Ni(II) in the feed solution were increased to 15 mg/L, and reactors were operated for a further 26 days.

Materials and methods

Statistical analysis

Aerobic granular systems operation

Multivariate methods were applied to the mean values of three replicates from each treatment. To avoid any effects of scale, the initial variable values were standardized mean centered and autoscaled to variance prior to analysis. The relationships within different samples on the basis of optical densities of each well were determined by principal component analysis (PCA). Hierarchical cluster analysis (HCA) is a technique used for classifying objects into different groups. The clusters are formed by grouping objects according to similarity, and the results are presented in the form of dendrograms, which is convenient in visualizing the distances between objects. Among the various ways, the between-groups linkage technique was adopted to cluster data, which defines the distance between two clusters as the average of all the pairs of distances between elements of both clusters. Similarities and dissimilarities were quantified by Square Euclidean distance measurements (Howard 1999). PCA, HCA, and correlation analysis were performed with Statistical Package for the Social Sciences (SPSS; version 13.0, SPSS Inc.).

Aerobic granules seeded in this study were cultivated in a 4-L SBR (100 cm in height and 8 cm in diameter) at room temperature 20±3 °C. The reactor was operated sequentially in 4-h cycles with 4 min of feeding, 20 min of anaerobic period, 210 min of aeration, 1 min of settling, and 5 min of effluent discharge. Effluent was discharged at the middle port of the reactor. Fine air bubbles for aeration were supplied through a fine-bubble diffuser at the reactor bottom with an airflow rate of 2.5 cm/s. Synthetic wastewater with the following composition was used in all reactors: glucose 100 mg/L, sodium acetate 120 mg/L, dissoluble amylum 125 mg/L, peptone 15 mg/L, meat extract 10 mg/L, NH4Cl 100 mg/L, KH2PO4 7.5 mg/L, Na2HPO4∙3H2O 20 mg/L, NaHCO3 200 mg/L, CaCl2 15 mg/L, MgSO4∙7H2O 12.5 mg/L, and trace solution 10 mL/L. The composition of the trace elements solution was CoCl2∙6H2O 0.25 g/L, MnSO4∙7H2O 2.5 g/L, and ZnCl2 0.2 g/L. The reactor was operated without pH control and varied between 7.0 and 8.5. After 1-year operation, aerobic granules were taken out and equally distributed into three identical SBRs (R1–R3) of the same geometrical configuration and operation conditions with the parent reactor, resulting in an initial biomass concentration of 4.0 g SS/L in each reactor. Seeded aerobic granules had granule size of 4–10 mm and mean

Biolog method Biolog Ecoplate was used to analyze the substrate utilization patterns of various sludge microbial communities. Sludge (4 g field-moist weight) was shaken with 36 mL of autoclave-sterilized saline solution (0.85% NaCl, w/v) for 2 h, and the suspension was settled for 20 min. The initial optical density (OD590) of super-cell suspension was 0.05, and a 150-µL aliquot supernatant was inoculated into each Ecoplate well. Following 12, 24, 48, 72, and 96 h of incubation at 28 °C, absorbance data were recorded at 590 nm with an ELISA Reader (biology microbial identification systems, BIOLOG Co., Ltd., Hayward, CA, USA).

Analytical methods Biomass concentration, COD, and N concentrations were measured according to Standard Methods (APHA 1998). The heat extraction method reported by Li and Yang (2007) was modified to extract extracellular polymeric substances

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(EPS) from the aerobic granules in this study. Aerobic granules were first dewatered by centrifugation in a 20-mL tube at 4,000 rpm for 5 min. The sludge pellet in the tube was then resuspended in 0.05% NaCl solution to its original volume of 20 mL. The sludge suspension was heated to 60 °C in a water bath for 30 min, and the sludge mixture was then centrifuged at 8,000 rpm for 30 min. The supernatant was collected and used as the EPS extraction of the sludge. EPS was normalized as the sum of proteins and polysaccharides, which were analyzed using modified Lowry method (Frølund et al. 1996) and anthrone method (Raunkjaer et al. 1994), respectively. DA assays were determined with a standard examination method (Japan Association for Sewage 1984) with the following procedure: (1) 0.5 mL TTC (4 g/L) and 2 mL Tris–HCl buffer solution (0.1 mol/L) were added to 2 mL sludge (4 g field-moist weight sludge was mixed with 36 mL of autoclavesterilized 0.85% saline solution for 3 h) and 0.5 mL supernatant; (2) the mixed solution was incubated at 37±1 °C for 2 h in the dark; (3) two drops of concentrated sulfuric acid were added to stop the deoxidization reaction in test tubes, and 5 mL methanol was added to extract the color product of TF from cells; and (4) the reaction mixture was centrifuged at 4,000 rpm for 5 min, and the absorbance of the supernatant was checked at 486 nm with a UV spectrometer. Values of DA are expressed as mg TF/g SS·h and are presented as an average of three replicates.

Results SBR performance After the addition of metal ions, aerobic granular systems present a great variance in biomass concentration towards the different metals studied (Fig. 1). In the control system without metal ions addition, the biomass concentration increased slowly and gradually from 5.0 to 8.6 g/L until the end of the experiment. While in the Cu(II)-fed system, an obvious biomass reduction was observed all along this experiment probably due to Cu(II) toxicity on the microorganisms. There is now a considerable amount of evidence documenting a decrease in the microbial biomass as a result of long-term exposure to heavy metals contamination such as Cu(II), Cr(II), and Cd(II) (Yao et al. 2003; Sheng et al. 2005; Lee et al. 2009). However, as compared with the control system, even though Ni(II) concentration suddenly increased from 5 to 15 mg/L on day 27, biomass concentration showed a continual and more rapid upward trend in the Ni(II)-fed system, except for a sudden decrease on day 40 due to the biomass taken out to keep the reactor from being blocked. From the biomass concentration profiles toward different metal ions, this study indicated

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that a Cu(II) concentration of 5 mg/L was high enough to inhibit the biomass growth of aerobic granules, while Ni(II) could stimulate the biomass yield under its concentration as high as 15 mg/L. In biological wastewater treatment systems, the most direct manifestation can be revealed by the treatment performance in COD biodegradation and nitrification (Fig. 2). The control system showed excellent and stable treatment performance with respective average COD and NH4+–N removal of 92.4% and 98.7%. The effects of Cu (II) addition on the performance of Cu(II)-fed system were significant. Five milligrams per liter Cu(II) slightly inhibited COD and NH4+–N removal during the first 2 weeks, but caused evident NO2−–N accumulation. These phenomena implied that nitrite-oxidizing bacteria could be the most sensitive to the toxic effect of Cu(II) from among the primary microbial components. After that, due to the gradually reduced biomass and further addition of Cu(II) to 15 mg/L, SBR performance was obviously deteriorated with COD removal between 60% and 80% and NH4+–N removal around 20%. Nevertheless, Ni(II) has a milder toxic effect on bioactivity of aerobic granules in contrast to Cu(II). The treatment performance of Ni(II)-fed system was not affected at 5 mg/L Ni(II), whereas the addition of 15 mg/L Ni(II) caused a slight reduction in COD and NH4+–N removal. Contrary to Cu(II), Ni(II) addition had no significant inhibition to nitrite oxidation. The higher sensitivity of ammonia oxidation versus nitrite oxidation to the presence of Ni(II) is consistent with an earlier observation of nitrification inhibition (Hu et al. 2002, 2004). It should be pointed out that due to the higher sensitivity of nitrifying bacteria versus heterotrophs (Principi et al. 2006; Sirianuntapiboon and Boonchupleing 2009), nitrification inhibition was more apparent in both heavy metal-fed systems.

Fig. 1 Time profile of biomass concentration in reactors

1970 Fig. 2 Treatment performance in sequencing batch reactors. a, b COD concentration and removal efficiency, respectively. c, d NH4+–N concentration and removal efficiency, respectively

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a

b

c

d

e

f

Extracellular polymeric substances EPS are biopolymers consisting of proteins, polysaccharides, nucleic acid, humic substances, etc. Often the composition and quantity of the EPS will vary depending on the type of microorganisms, age of the aerobic granules, and the different environmental conditions under which the aerobic granules exist. Microorganisms are known to regulate EPS synthesis and modify EPS properties as a microbial response against the effects of toxic chemicals (Allison et al. 2000; Jiang et al. 2004b). The effects of Cu (II) and Ni(II) on production of EPS in aerobic granules were exhibited in Fig. 3. In contrast with the control system, at Cu(II) concentrations of 5 and 15 mg/L and Ni (II) concentration of 5 mg/L, the content of EPS changed slightly from 34.2 to 40.8 mg/g SS, while it increased sharply from 34.5 to 55.4 mg/g SS when Ni(II) concentration was increased from 5 to 15 mg/L. This increment was mainly attributed to the production of proteins. In other words, the production of proteins was greatly stimulated only at Ni(II) concentration of 15 mg/L.

Fig. 3 Extracellular polymeric substances under different metal ion concentrations (26th day, 5 mg/L; 52nd day, 15 mg/L)

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422.08 mg TF/g SS·h, and since then, it fluctuated between 26.38 and 281.73 mg TF/g SS·h, which was fairly low as compared with DA values in the control and Ni(II)-fed systems. Similar to the conclusion obtained from biomass concentration profiles, 5 mg/L Cu(II) inhibited microbial viability in aerobic granules severely. In both control and Ni(II)-fed systems, DA profiles fluctuated widely in a similar trend with respective mean values of 1,919.01 and 2,032.69 mg TF/g SS·h. Although the exact reason is unclear, this phenomenon may be ascribed to temperature changes existing in the experimental period. In this case, it can be concluded that, even at a concentration of 15 mg/L, Ni(II) still stimulated microbial viability of aerobic granules to some extent. Fig. 4 Time profile of dehydrogenase activity of aerobic granules

Dehydrogenase activity Dehydrogenase enzyme, as an intracellular enzyme, is affected by the number of microbes and bacteriostat agents existing in environments. It can be used as a valuable indicator to assess the microbial viability in environments and has been justified as an ecotoxicological bioassay (Barrena et al. 2008). DA is an indicator of the primary activity of microorganisms, since it is related to the cellular respiratory processes, whether this activity is aerobic or not. Figure 4 shows the variation of DA in all of the treatments. Just after 2 days of the addition of 5 mg/L Cu(II), the value of DA was extremely decreased from 2,005.26 to

a

Microbial community biodiversity PCA (Fig. 5a) and HCA (Fig. 5b), using all 31 carbon sources, revealed more intuitively the patterns of sludge heterotrophic community diversity and offered robust discrimination among different sludge samples. The Cu(II) and Ni(II) treatments were clearly clustered and distant from the heavy metal-free control, which indicated that the addition of Cu(II) and Ni(II) significantly affected microbial community structure and biodiversity of aerobic granules. The first three PCs (PC1, PC2, and PC3) accounted for 60.23, 11.53, and 9.19%, respectively, of the total variation in the Biolog data, for a cumulative total of 80.95%. The projection of each sludge sample (in its respective cluster) onto each substrate vector in the PCA

b

Fig. 5 Classification of Biolog profiles by (a) principal component analysis and (b) hierarchical cluster analysis of substrate utilization patterns for aerobic granules

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confirmed that clusters A (metal-free control) and C (5 mg/L Ni(II)-fed) had higher utilization values for most carbon sources, and analysis of PC1 even indicates that cluster C utilized a number of carbon sources loaded on PC1 to a relatively greater degree than cluster A. Both samples B (5 mg/L Cu(II)-fed) and E (15 mg/L Ni(II)-fed) formed a joint cluster with PC1 and PC3 loadings close to sample A. On the basis of PC1 and PC3 analysis, cluster D (15 mg/L Cu (II)-fed) showed a relatively lesser utilization of most carbon sources, which means a significant reduced biodiversity of aerobic granules at 15 mg/L Cu(II). With HCA, the singlelinkage dendrogram was given. A1–A3, B1–B3, C1–C3, D1–D3, and E1–E3 formed a compact cluster, respectively. Cluster E was the closest to B (distance 1.83±0.04) and formed a group; at the same time, A was close to C (distance 2.18±0.18) and formed another group. After that, groups A and C joined groups B and E, while group D is separated from the other groups, which is similar to the PCA graph. The above-mentioned findings demonstrated that the addition of Cu(II) (especially at a concentration of 15 mg/L) drastically reduced the biodiversity of aerobic granules, while the toxic effect of Ni(II) addition was milder.

Discussion Like all the essential elements, Cu(II) and Ni(II) often stimulate the microbial growth, at relatively low concentrations; however, both metals are toxic at relatively high concentrations. The most widely accepted theory to explain the toxic or inhibitory effect is the inactivation or damage of some enzymes within the cell (McCarthy 1964). However, contrary to many other metal ions including Ni (II), the major mechanism of Cu(II) microbial toxicity is not intracellular inactivation but cytoplasmic membrane disruption (Sani et al. 2001; Hu et al. 2003). As a unique redoxactive metal, Cu(II) has been reported to induce cytotoxicity by producing free hydroxyl radical via Feton-type reactions and promoting membrane lipid peroxidation (Howlett and Avery 1997; Stillman and Presta 2000). Cu(II) sorption on biomass is strong and rapid, which in turn caused a rapid loss of membrane integrity (Hu et al. 2004). Thus, as depicted in this study, Cu(II) appeared to be more toxic than Ni(II) on microorganisms of aerobic granules. The addition of Cu (II) even at a concentration of 5 mg/L drastically reduced the biomass concentration, bioactivity, and biodiversity of aerobic granules, and certainly deteriorated reactor performance in COD and NH4+–N removal. In fact, it was often found that Cu(II) has much stronger toxic effects than many other heavy metals like Ni(II), Zn(II), Pb(II), Cr(III), etc. (Lee et al. 1997; Juliastuti et al. 2003; Colussi et al. 2009). It is worthwhile to note that SBR performance was not completely inhibited even at 15 mg/L Cu(II)

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(Fig. 2). However, it was demonstrated that a Cu(II) concentration of 1.2 mg/L would exhibit complete inhibition on the nitrification process of activated sludge (Juliastuti et al. 2003). That is, according to this study, the aerobic granular system was also susceptible to Cu(II) toxicity due to the unique mode of action of Cu(II), and meanwhile it exhibited some advantages in resisting this toxicity. These advantages are highlighted in Ni(II)-fed aerobic granular system. The compact-structured aerobic granules provided microorganisms with the ability to resist Ni(II) toxicity. Even at a concentration of 15 mg/L, Ni(II) has no obvious toxic effects on aerobic granular system, and it even stimulated the biomass yield and bioactivity to some extent. However, activated sludge suspended growth systems showed a lower tolerance to Ni(II) toxicity. Research carried out by Sujarittanonta and Sherrard (1981), who worked with activated sludge growing in a continuous bioreactor setup, indicates that addition of 1 or 5 mg/L Ni(II) enhanced the maximum biomass yield of activated sludge, while the nitrification process was significantly suppressed even by the addition of 1 mg/L Ni(II). In acclimated activated sludge systems, Yetis and Gokcay (1989) have reported that the substrate removal efficiency was not adversely affected by the presence of Ni (II) up to a concentration of 10 mg/L, while 5 mg/L Ni(II) had some stimulatory effects. On the other hand, a number of researchers have reported only inhibition to the growth of activated sludge, due to the presence of Ni(II), as opposed to the enhancement of growth at small nickel concentrations. Research performed by McDermott et al. (1965) indicated that a continuous mode activated sludge plant was able to withstand the addition of 1 mg/L Ni(II) in the feed solution; however, addition of 2.5, 5, or 10 mg/L of Ni(II) resulted to up to a 5% reduction of the biochemical oxygen demand removal efficiency. Ong et al. (2004) operated an SBR under similar conditions with our study for activated sludge growth, reporting 23% reduction in the specific oxygen uptake rate with the addition of 5 mg/L Ni (II) in the influent, whereas 10 mg/L Ni(II) addition significantly affected the SBR performance in terms of biomass concentration and TOC removal efficiency. The elevated level of toxicity tolerance in aerobic granular system could be due to the concentration gradient developed within granules by diffusional resistance, which can protect the microorganisms by reducing the concentration of the chemicals below some threshold value to avoid inhibition. The tolerance of biological treatment systems to heavy metals can be enhanced greatly by acclimation (Özbelge et al. 2007). Hereby, aerobic granules could provide a buffer for microorganisms to acclimate higher heavy metals concentration such as by alternating biochemical pathways, which allow cells to continue growing (Nies 1992)

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or by redistributing bacteria to best meet the needs of each other and the community (Watnick and Kolter 2000). Bacteria with a good acclimation for heavy metals will distribute themselves near the granule surface, where heavy metals concentrations are higher. It has been reported that anaerobic or aerobic granules could provide microbial cells with the ability to resist aromatic chemicals such as phenol and p-nitrophenol (Fang and Chan 1997; Jiang et al. 2004a; Yi et al. 2006). Compared with activated sludge flocs, which would have a lower diffusional resistance, a strongly packed microbial structure, in the form of either biofilm or granules, would have a higher tolerance level towards toxic chemicals. In this study, aerobic granules were also demonstrated to protect microorganisms against the toxic heavy metals. It should be pointed out that aerobic granules would also have a high diffusion resistance to the substrate and oxygen (Chiu et al. 2007; Li et al. 2008). Diffusion limitation problems often existing in large granules can reduce the substrate removal rate and also potentially diminish granule stability (Tay et al. 2002; Li and Liu 2005). Thus, an optimal granule size range should be taken into account for effective and stable operation for large-scale SBR (Toh et al. 2003). In addition to the inhibition of diffusion, it was also proposed that the elevated level of Ni(II) toxicity tolerance could be due to the increased biomass concentration and EPS content at 15 mg/L Ni(II)-fed system as mentioned before. As biomass concentration increased, the toxic effects of heavy metals would be offset or reduced by the biomass due to higher metal sorption to the biomass and low heavy metals/biomass ratios (Kim et al. 2006; Delgado et al. 2010). Meanwhile, the EPS matrix could function as a protective barrier for the bacteria to chemicals that it has the potential to interact with, such as positively charged compounds (Henriques and Love 2007). Heavy metal ions can form a complex with EPS produced by the biomass, which in turn would reduce the available free metal ions in the bulk media. As the toxicity of metals in the environment is related to the presence of free metal ion concentrations rather than the total or complexed metal concentration (Campbell 1995), the increase in biomass concentration and EPS would reduce the toxicity effect of heavy metals toward the biomass activity. Meanwhile, the EPS matrix could form a protective shield for the cells against the toxic substances from reaching the microorganisms within the aerobic granules by acting as a diffusion limitation barrier (Wingender et al. 1999). The EPS barriers appear to function through sorption of matrix components with the heavy metals, as well as through retardation of toxin penetration. Microorganisms are also known to regulate EPS synthesis and modify EPS properties as a microbial response against the effects of toxic chemicals (Allison et al. 2000). According to the results from the present work, it can be speculated that the

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EPS production was closely related to the extent of toxic effect of metal ions. At Cu(II) concentrations of 5 and 15 mg/L, the EPS production was not promoted because microbial metabolism activity was significantly inhibited, while the toxic effect of 5 mg/L Ni(II) might be counteracted by the increased biomass concentration, and there was no need for microorganisms to regulate the EPS synthesis. Further addition of 15 mg/L Ni(II) affected SBR performance slightly, and microorganisms secreted more EPS (especially proteins) to protect themselves. Sheng et al. (2005) have proved that most EPS were produced at a certain concentration of heavy metals, whereas heavy metals at a high concentration reduced the production of EPS. As a microbial response to other toxic compounds such as phenol, Jiang et al. (2004b) also found that the content of EPS in aerobic granules was similar under phenol loading of 1, 1.5, and 2 kg/m3 day, but an increase in phenol loading to 2.5 kg/m3 day was associated with 216% and 33% increase in the content of proteins and polysaccharides, respectively. The preferential production of proteins over polysaccharides in EPS has also been observed in aerobic granules and biofilms exposed to phenol (Fang et al. 2002). One possible explanation for the elevated production of proteins is the induction of heat shock-like proteins as a defense mechanism against high heavy metal ions concentrations (Benndorf et al. 2001). In summary, this study highlights the advantage of aerobic granules to protect microorganisms against the toxic heavy metals, due to the concentration gradient developed within granules by diffusional resistance, increased biomass concentration, and promoted EPS production. Future research should be implemented on metal toxicity mechanisms against microorganisms and long-term toxic effect on stability and structure of aerobic granules. Acknowledgements This work was supported by the China Postdoctoral Science Foundation (No. 20090461216), the Postdoctoral Innovation Foundation of Shandong Province (No. 200903049), and the Independent Innovation Foundation of Shandong University (No. 2009TS022).

References Adav SS, Lee DJ, Lai JY (2008) Intergeneric coaggregation of strains isolated from phenol-degrading aerobic granules. Appl Microbiol Biotechnol 79:657–661 Adav SS, Lee DJ, Lai JY (2009) Biological nitrification–denitrification with alternating oxic and anoxic operations using aerobic granules. Appl Microbiol Biotechnol 84:1181–1189 Allison DG, Maira-Litran T, Gilbert P (2000) Antimicrobial resistance of biofilms. In: Evans LV (ed) Biofilms: recent advances in their study and control. Harwood, Amsterdam, pp 149–166 APHA (1998) Standard methods for the examination of water and wastewater, 20th edn. American Public Health Association, Washington

1974 Barrena R, Vazquez F, Sanchez A (2008) Dehydrogenase activity as a method for monitoring the composting process. Bioresour Technol 99:905–908 Benndorf D, Loffhagen N, Babel W (2001) Protein synthesis patterns in Acinetobacter calcoaceticus induced by phenol and catechol show specificities of responses to chemostress. FEMS Microbiol Lett 200:247–252 Beun JJ, van Loosdrecht MCM, Heijnen JJ (2002) Aerobic granulation in a sequencing batch airlift reactor. Water Res 36:702–712 Campbell PGC (1995) Interaction between trace metals and aquatic organisms: a critique of the free-ion activity model. In: Tessier A, Turner DR (eds) Metal speciation and bioavailability in aquatic systems. Wiley, NY, pp 45–102 Chiu ZC, Chen MY, Lee DJ, Wang CH, Lai JY (2007) Oxygen diffusion in active layer of aerobic granule with step change in surrounding oxygen levels. Water Res 41:884–892 Colussi I, Cortesi A, Vedova LD, Gallo V, Robles FKC (2009) Startup procedures and analysis of heavy metals inhibition on methanogenic activity in EGSB reactor. Bioresour Technol 100:6290–6294 de Kreuk MK, Heijnen JJ, Van Loosdrecht MCM (2005) Simultaneous COD, nitrogen, and phosphate removal by aerobic granular sludge. Biotechnol Bioeng 90:761–769 Delgado LF, Schetrite S, Gonzalez C, Albasi C (2010) Effect of cytostatic drugs on microbial behaviour in membrane bioreactor system. Bioresour Technol 101:527–536 Fang HHP, Chan OC (1997) Toxicity of phenol towards anaerobic biogranules. Water Res 31:2229–2242 Fang HHP, Xu LC, Chan KY (2002) Effects of toxic metals and chemicals on biofilm and biocorrosion. Water Res 36:4709–4716 Frølund B, Palmgren R, Keiding K, Nielsen PH (1996) Extraction of extracellular polymers from activated sludge using a cation exchange resin. Water Res 30:1749–1758 Henriques IDS, Love NG (2007) The role of extracellular polymeric substances in the toxicity response of activated sludge bacteria to chemical toxins. Water Res 41:4177–4185 Howard PJA (1999) Analysis of inter-sample distances from BIOLOG plate data in Euclidean and simplex spaces. Soil Biol Biochem 31:1323–1330 Howlett NG, Avery SV (1997) Induction of lipid peroxidation during heavy metal stress in Saccharomyces cerevisiae and influence of plasma membrane fatty acid unsaturation. Appl Environ Microbiol 63:2971–2976 Hu Z, Chandran K, Grasso D, Smets BF (2002) Effect of nickel and cadmium speciation on nitrification inhibition. Environ Sci Technol 36:3074–3078 Hu Z, Chandran K, Grasso D, Smets BF (2003) Impact of metal sorption and internalization on nitrification inhibition. Environ Sci Technol 37:728–734 Hu Z, Chandran K, Grasso D, Smets BF (2004) Comparison of nitrification inhibition by metals in batch and continuous flow reactors. Water Res 38:3949–3959 Jiang HL, Tay JH, Maszenan AM, Tay STL (2004a) Bacterial diversity and function of aerobic granules engineered in a sequencing batch reactor for phenol degradation. Appl Environ Microbiol 70:6767–6775 Jiang HL, Tay JH, Tay STL (2004b) Changes in structure, activity and metabolism of aerobic granules as a microbial response to high phenol loading. Appl Microbiol Biotechnol 63:602–608 Juang YC, Adav SS, Lee DJ, Lai JY (2009) Biodiversity in aerobic granule membrane bioreactor at high organic loading rates. Appl Microbiol Biotechnol 85:383–388 Juliastuti SR, Baeyens J, Creemers C, Bixio D, Lodewyckx E (2003) The inhibitory effects of heavy metals and organic compounds on the net maximum specific growth rate of the autotrophic biomass in activated sludge. J Hazard Mater 100:271–283

Appl Microbiol Biotechnol (2010) 86:1967–1975 Kim KT, Kim IS, Hwang SH, Kim SD (2006) Estimating the combined effects of copper and phenol to nitrifying bacteria in wastewater treatment plants. Water Res 40:561–568 Lee YW, Ong SK, Sato C (1997) Effects of heavy metals on nitrifying bacteria. Water Sci Technol 36:69–74 Lee YW, Tian Q, Ong S, Sato C, Chung J (2009) Inhibitory effects of copper on nitrifying bacteria in suspended and attached growth reactors. Water Air Soil Pollut 203:17–27 Li Y, Liu Y (2005) Diffusion of substrate and oxygen in aerobic granule. Biochem Eng J 27:45–52 Li XY, Yang SF (2007) Influence of loosely bound extracellular polymeric substances (EPS) on the flocculation, sedimentation and dewaterability of activated sludge. Water Res 41:1022–1030 Li Y, Liu Y, Shen L, Chen F (2008) DO diffusion profile in aerobic granule and its microbiological implications. Enzyme Microb Technol 43:349–354 Liu Y, Wang ZW, Tay JH (2005) A unified theory for upscaling aerobic granular sludge sequencing batch reactors. Biotechnol Adv 23:335–344 McCarthy PL (1964) Anaerobic waste treatment fundamentals. Part III: toxic materials and their control. Public Works 95:91–94 McDermott GN, Post MA, Jackson BN, Ettinger MB (1965) Nickel in relation to activated sludge and anaerobic digestion processes. J Water Pollut Control Fed 37:163–177 Ni BJ, Xie WM, Liu SG, Yu HQ, Wang YZ, Wang G, Dai XL (2009) Granulation of activated sludge in a pilot-scale sequencing batch reactor for the treatment of low-strength municipal wastewater. Water Res 43:751–761 Nicolau A, Martins MJ, Mota M, Lima N (2005) Effect of copper in the protistan community of activated sludge. Chemosphere 58:605–614 Nies DH (1992) Resistance to cadmium, cobalt, zinc, and nickel in microbes. Plasmid 27:17–28 Ong SA, Toorisaka E, Hirata M, Hano T (2004) Effects of nickel(II) addition on the activity of activated sludge microorganisms and activated sludge process. J Hazard Mater 113:111–121 Özbelge TA, Özbelge HÖ, Altınten P (2007) Effect of acclimatization of microorganisms to heavy metals on the performance of activated sludge process. J Hazard Mater 142:332–339 Principi P, Villa F, Bernasconi M, Zanardini E (2006) Metal toxicity in municipal wastewater activated sludge investigated by multivariate analysis and in situ hybridization. Water Res 40:99–106 Raunkjaer K, Hvitved-Jacobsen T, Nielsen PH (1994) Measurement of pools of protein, carbohydrate and lipid in domestic wastewater. Water Res 28:251–262 Sani RK, Peyton BM, Brown LT (2001) Copper-induced inhibition of growth of Desulfovibrio desulfuricans G20: assessment of its toxicity and correlation with those of zinc and lead. Appl Environ Microbiol 67:4765–4772 Japan Association for Sewage (1984) Standard examination method for wastewater. Japan Association for Sewage, Japan, pp 299– 300 Sheng GP, Yu HQ, Yue ZB (2005) Production of extracellular polymeric substances from rhodopseudomonas acidophila in the presence of toxic substances. Appl Microbiol Biotechnol 69:216–222 Sirianuntapiboon S, Boonchupleing M (2009) Effect of bio-sludge concentration on the efficiency of sequencing batch reactor (SBR) system to treat wastewater containing Pb2+and Ni2+. J Hazard Mater 166:356–364 Stillman MJ, Presta A (2000) In molecular biology and toxicology of metals, edn. Taylor & Francis, New York Sujarittanonta S, Sherrard JH (1981) Activated sludge nickel toxicity studies. J Water Pollut Control Fed 53:1314–1322 Tay STL, Ivanov V, Yi S, Zhuang WQ, Tay JH (2002) Presence of anaerobic bacteroides in aerobically grown microbial granules. Microb Ecol 44:278–285

Appl Microbiol Biotechnol (2010) 86:1967–1975 Toh SK, Tay JH, Moy BYP, Ivanov V, Tay STL (2003) Size-effect on the physical characteristics of the aerobic granule in a SBR. Appl Microbiol Biotechnol 60:687–695 Vandevivere P, Ficara E, Terras C, Julies E, Verstraete W (1998) Copper-mediated selective removal of nitrification inhibitors from industrial wastewaters. Environ Sci Technol 32:1000–1006 Wang XH, Zhang HM, Yang FL, Xia LP, Gao MM (2007) Improved stability and performance of aerobic granules under stepwise increased selection pressure. Enzyme Microb Technol 41:205–211 Wang XH, Zhang HM, Yang FL, Wang YF, Gao MM (2008) Longterm storage and subsequent reactivation of aerobic granules. Bioresour Technol 99:8304–8309 Watnick P, Kolter R (2000) Biofilm, city of microbes. J Bacteriol 182:2675–2679 Wingender J, Neu TR, Flemming HC (1999) What are bacterial extracellular polymeric substances. In: Wingender J, Neu TR,

1975 Flemming HC (eds) Microbial extracellular polymeric substances: characterization, structure and function. Springer, Berlin, pp 1–19 Yao H, Xu J, Huang C (2003) Substrate utilization pattern, biomass and activity of microbial communities in a sequence of heavy metal-polluted paddy soils. Geoderma 115:139–148 Yetis U, Gokcay CF (1989) Effect of nickel(II) on activated sludge. Water Res 23:1003–1007 Yi S, Zhuang WQ, Wu B, Tay STL, Tay JH (2006) Biodegradation of p-nitrophenol by aerobic granules in a sequencing batch reactor. Environ Sci Technol 40:2396–2401 You SJ, Tsai YP, Huang RY (2009) Effect of heavy metals on nitrification performance in different activated sludge processes. J Hazard Mater 165:987–994 Zhang LL, Chen JM, Fang F (2008) Biodegradation of methyl t -butyl ether by aerobic granules under a cosubstrate condition. Appl Microbiol Biotechnol 78:543–550