Journal of Hazardous Materials 295 (2015) 153–160
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Aerobic granulation strategy for bioaugmentation of a sequencing batch reactor (SBR) treating high strength pyridine wastewater Xiaodong Liu a , Yan Chen a , Xin Zhang a,b , Xinbai Jiang a , Shijing Wu a , Jinyou Shen a,∗ , Xiuyun Sun a , Jiansheng Li a , Lude Lu a , Lianjun Wang a,∗∗ a Jiangsu Key Laboratory for Chemical Pollution Control and Resources Reuse, School of Environmental and Biological Engineering, Nanjing University of Science and Technology, Nanjing 210094, Jiangsu Province, China b Suzhou Institute of Architectural Design Co., Ltd, Suzhou 215021, Jiangsu Province, China
g r a p h i c a l
a r t i c l e
a b s t r a c t
i n f o
Article history: Received 26 August 2014 Received in revised form 6 March 2015 Accepted 9 April 2015 Available online 13 April 2015 Keywords: Aerobic granule Bioaugmentation Biodegradation High-throughput sequencing
a b s t r a c t Aerobic granules were successfully cultivated in a sequencing batch reactor (SBR), using a single bacterial strain Rhizobium sp. NJUST18 as the inoculum. NJUST18 presented as both a good pyridine degrader and an efficient autoaggregator. Stable granules with diameter of 0.5–1 mm, sludge volume index of 25.6 ± 3.6 mL g−1 and settling velocity of 37.2 ± 2.7 m h−1 , were formed in SBR following 120-day cultivation. These granules exhibited excellent pyridine degradation performance, with maximum volumetric degradation rate (Vmax ) varied between 1164.5 mg L−1 h−1 and 1867.4 mg L−1 h−1 . High-throughput sequencing analysis exhibited a large shift in microbial community structure, since the SBR was operated under open condition. Paracoccus and Comamonas were found to be the most predominant species in the aerobic granule system after the system had stabilized. The initially inoculated Rhizobium sp. lost its dominance during aerobic granulation. However, the inoculation of Rhizobium sp. played a key role in the startup process of this bioaugmentation system. This study demonstrated that, in addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula is equally important for the formation of the aerobic granule. © 2015 Elsevier B.V. All rights reserved.
∗ Corresponding author. Tel. +86 25 84303965; fax: +86 25 84303965. ∗∗ Corresponding author. Tel. +86 25 84315941; fax: +86 25 84315941. E-mail addresses: [email protected] (J. Shen), [email protected] (L. Wang). http://dx.doi.org/10.1016/j.jhazmat.2015.04.025 0304-3894/© 2015 Elsevier B.V. All rights reserved.
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1. Introduction Aerobic granules have been employed in treating toxic and recalcitrant organics, such as phenol, p-nitrophenol and 2,4dichlorophenol, exhibiting high resistance to these toxic and recalcitrant compounds [1]. The high toxic resistance and degradation ability of the aerobic granules are partially attributed to the unique structure of granules, where the dense microbial cells and extracellular polymeric substances (EPS) set a barrier for mass transfer and lower the concentration of toxics on the inner cells [2]. Considering the compact structure, large microbial density and high toxic resistance, aerobic granules might be the core of bioaugmentation [1]. In the formation of aerobic granules, hydraulic selection pressure to select and retain compact microbial aggregates plays a key role. In addition, cell aggregation including autoaggregative or coaggregative interactions may be promoted for aerobic granulation. In the previous studies, aerobic granulation from coaggregative or autoaggregative phenol-degrading strains has been reported for enhanced phenol biodegradation [3–5]. However, detailed information regarding the roles that coaggregation/autoaggregation played in aerobic granulation has not yet been fully investigated. Although, several works have been reported to isolate efficient microbial species for pyridine biodegradation [6], most of the work reported on this heterocyclic pollutant degradation pertained to flask level experiments. Only limited reports with technological orientation are available by now [7]. Reports on aerobic granulation aiming at the bioremediation of pyridine waste are rare [8]. In addition, formation of pyridine-degrading aerobic granules from pure cultures capable of degrading pyridine has not previously been explored to our knowledge. Therefore, the present study aimed to investigate the feasibility of aerobic granulation from Rhizobium sp. NJUST18, which was both a functional pyridine degrader and a good autoaggregator. The effectiveness of the aerobic granules in treating high strength pyridine wastewater was evaluated, and the bacterial community structure of the aerobic granules was also included. 2. Materials and methods
bioreactor was operated in a thermostatic bath to keep cultivation temperature stable at 30 ◦ C. The SBR was initially seeded with 2 g (measured in dry weight) Rhizobium sp. NJUST18 pure culture and fed with the synthetic influent. After inoculation, the SBR was operated in open condition, with the influent and the air introduced without sterilization. Operational parameters for different phases of the cultivation procedure are presented in Table 1. During the start-up period, i.e., stage I (1–35 days), the bioreactor was operated at a cycle of 24 h:5 min of feeding without stirring, 23.5 h of aerobic reaction, 12 min of settling, 3 min of effluent withdrawal and 10 min of idling. For the first week, in addition to 1000 mg L−1 pyridine, 1000 mg L−1 acetate sodium was added into synthetic influent as supplementary substrate for biomass growth stimulation. One week later, acetate sodium in synthetic influent was reduced progressively, and finally reduced to 0 mg L−1 on day 30. Thereafter, influent pyridine concentration increased to 4000 mg L−1 step by step, while the cycle time and settling time progressively decreased to 6 h and 2 min, respectively. 120 days later, mature granules were observed in SBR, with stable operation of the SBR achieved. In order to investigate the effect of pyridine concentration on pyridine biodegradation in the bioaugmented SBR, 1000–8000 mg L−1 pyridine was added into the synthetic influent to make initial pyridine in SBR over the concentration range between 500 and 4000 mg L−1 . In order to reveal the key role of Rhizobium sp. NJUST18 during the startup of the bioaugmented SBR, a SBR inoculated with common activated sludge was operated as a control. The common activated sludge used was taken from the secondary sedimentation tank of a municipal sewage treatment plant located in Nanjing. The control SBR was initially seeded with 6 g activated sludge (measured in dry weight) and was operated under the same conditions as the bioaugmented SBR, except for the inocula used. Specifically, 1000 mg L−1 pyridine and 1000 mg L−1 acetate sodium was added into the synthetic influent as the carbon sources for the first week. One week later, acetate sodium in synthetic influent was reduced to 800 mg L−1 while pyridine remained the same as the first week. However, due to the complete failure, the control SBR stopped running after 2 weeks.
2.1. Microorganism and synthetic wastewater The isolation and cultivation of Rhizobium sp. NJUST18 was described in our previous study [9]. The 16S rRNA sequence (comprising 1381 nucleotides) was deposited in the GenBank database under accession no. JN106368. The synthetic influent of the sequential batch reactor (SBR) was as follows: Na2 HPO4 • 12H2 O (3.057 g L−1 ), KH2 PO4 (0.743 g L−1 ), NH4 Cl (1 g L−1 ), MgSO4 • 7H2 O (0.89 g L−1 ), KCl (0.35 g L−1 ), CaCl2 (0.20 g L−1 ), FeCl3 (0.03 g L−1 ), pyridine and sodium acetate at desired concentrations, and 10 mL L−1 trace element solution SL-4 [10]. 2.2. Aerobic granules cultivation The aerobic granules were cultivated in a column-type sequential batch reactor (SBR) with inner diameter of 6 cm, height of 100 cm and effective volume of 2.2 L. The sequential operation of the reactors was automatically controlled by the timers. During the filling phase, the synthetic influent was introduced through ports located at the bottom of the bioreactor. While, fine air bubbles for aeration were supplied by means of air bubble diffusers placed at the bottom of the SBR at volumetric flow rate of 0.2 m3 h−1 and superficial gas velocity of 0.02 m s−1 . The effluent was withdrawn through the outlet ports positioned at the medium height in this column reactor, resulting in volumetric exchange ratio of 50%. The
2.3. Analytical methods Pyridine was quantified by HPLC (Waters 2996, Waters Incorporation, USA) through authentic standard. Before analysis, water samples were passed through a 0.22 m filter. The HPLC analysis was conducted at room temperature using a Waters RP18 column (5 m, 4.6 × 250 mm) and a UV–vis detector. The mobile phase was a mixture of 70% methanol and 30% water pumped at a flow rate of 1.00 mL/min. The analysis was performed at 254 nm, with column temperature at 35 ◦ C. Measurements of the parameters, such as sludge volume index (SVI), mixed liquor suspended solid (MLSS) and mixed liquor volatile suspended solids (MLVSS), were carried out according to APHA standard methods [11]. The settling velocity of the sludge was measured by recording the traveling time of individual granules from the top to the bottom of a cylinder [12]. The granule size was manually measured using a fine-scale ruler [13]. Scanning electron microscopy (SEM) observation was carried out according to Ho et al. [4]. Transmission electron microscopy (TEM) observation was carried out according to Bhattacharya et al. [14]. The effluent samples for the determination of the residual TOC concentration were sampled 2 h after the complete exhaustion of the pyridine. TOC concentration was measured using a Germany Elementar vario TOC analyzer.
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Table 1 Operational parameters applied during different cultivation phases. Stage
I
II
III
IV
V
VI
Time (day) Influent pyridine concentration (mg L−1 ) Influent NaAC concentration (mg L−1 ) Cycle duration (h) Settling time (min) Aeration time (min) Pyridine loading rate (kg m−3 d−1 )
1–35 1000 1000–0 24 12 1410 0.5
36–45 1500 0 12 10 692 1.5
46–55 2000 0 12 5 697 2.0
56–65 2000 0 8 3 459 3.0
66–85 4000 0 6 3 339 8.0
86–120 4000 0 6 2 340 8.0
2.4. DNA extraction, PCR and High-throughput sequencing The microbial community structure of the aerobic granules in the SBR system was profiled using high-throughput sequencing technology. About 1 g sample of mature aerobic granules was
collected at the end of stage VI. DNA of the aerobic granules was extracted from the aerobic granule sample using the FastDNA® SPIN for soil kit (MP Biomedicals, CA, USA) according to the manufacturer’s instruction. The concentration and purity of genomic DNA were measured using NanoDrop® ND-1000 (NanoDrop Technolo-
Fig. 1. TEM of Rhizobium sp. NJUST18 (a), biomass 2 weeks after inoculation (b), biomass 14 weeks after inoculation (c), biomass18 weeks after inoculation (d), and SEM of mature granule (e and f).
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Fig. 2. The MLSS, MLVSS and settling time profiles during 120 days’ operation.
gies, Willmington, DE, USA) to ensure the sample concentration higher than 20ng/L and A260/A280 between 1.80 and 2.00. The extracted DNA was amplified by polymerase chain reaction (PCR) using TaKaRa Ex Taq® (TaKaRa Bio Japan) according to the manufacturer’s instruction. Primers 27F (5 -AGAGTTTGATYMTGGCTCAG-3 ) and 338R (5 TGCTGCCTCCCGTAGGAGT-3 ) which targeted V1/V2 hypervariable regions of bacterial 16S rRNA genes were selected. In order to tag the PCR products from each sample, specific 8 bases long identifier were included in both forward and reverse primers. For sequencing the PCR products, specific sequencing adaptors were attached to both primers. The extracted DNA was diluted to 20ng/L and used as the template DNA. Thermal cycling conditions were as follows: an initial denaturation at 98 ◦ C for 5 min, and 20 cycles at 98 ◦ C for 30 s, 50 ◦ C for 30 s, and 72 ◦ C for 40 s, with a final extension at 72 ◦ C for 10 min. Four parallel PCR products were pooled and purified by OMEGA E.Z.N.A Cycle-Pure Kit (Omega Bio-Tek Norcross, USA). The purity of purified PCR product was measured using NanoDrop® ND-1000 once again to ensure its A260/A280 between 1.70 and 1.90. The purified PCR product was quantified precisely using Qubit 2.0 Fluorometer (Life Technologies, Grand Island, USA). The purified library was diluted, denatured, rediluted, mixed with PhiX (equal to 30% of final DNA amount) as described in the Illumina library preparation protocols, and then applied to an Illumina Miseq system in Jiangsu Zhongyijinda Analytical & Testing Co., Ltd. for sequencing with the Reagent Kit v22 × 250 bp as described in the manufacture manual. DNA library building and data analysis were performed in according to Liang et al. [15] and Pala-Ozkok et al. [16]. 3. Results 3.1. Formation of pyridine aerobic degrading granule TEM image of Rhizobium sp. NJUST18 (Fig. 1a) showed that strain NJUST18 was a rod-shaped bacterium with flagella, which were assumed to correspond closely to the noted capability of autoaggregation. One week after the inoculation of the autoaggregative NJUST18, both biomass increase and obvious aggregation in SBR were observed. Two weeks later, the biomass in SBR showed fluffy, irregular, loose-structure morphology (Fig. 1b). Some small and white granules could be observed, however, these granules were easily broke up into small pieces if placed under vigorous shaking. 14 weeks after inoculation, with the increase of influent pyridine concentration, decrease of the cycle time and settling time, the flocs-like sludge gradually decreased and was replaced by small
Fig. 3. The settling velocities and SVI profiles during 120 days’ operation.
granules with diameter of 0.2–0.5 mm, while the color of the granules gradually changed from white to yellowish brown (Fig. 1c). In the following 4 weeks, the granules became denser and more regular in shape under high shear force which induced the biomass aggregates to secrete more exopolysaccharides [17,18]. Through the control of settling time, more flocculent sludge was washed out from the SBR, resulting in the accumulation of the aerobic granules with high settling velocity. After 18 weeks of inoculation, mature granules (Fig. 1d) with size of 0.5–1 mm were formed, leading to a stable operation of the SBR. The mature granules turned out to be smooth, with a solid surface (Fig. 1d and e). Through SEM observation, rod-shaped bacteria were found to be dominated at the granule surface (Fig. 1f), indicating that rod-shaped strains played an important role in the granule formation and granule stabilization. 3.2. Biomass profile and settling properties of pyridine aerobic degrading granules Since the settling time is a critical parameter and significantly influences the granulation process, a short settling time was generally considered favorable for the granule formation [2]. Fig. 2 shows the MLSS, MLVSS during the development period of the pyridine aerobic degrading granules. During the stage I, the settling time was fixed at 12 min. MLSS and MLVSS increased from initial 878.2 ± 30.8 mg L−1 and 853.6 ± 45.3 mg L−1 to 3582.0 ± 63.8 mg L−1 and 3451.1 ± 194.5 mg L−1 on day 30, respectively. Decreasing the settling time from 12 min to 10 min on stage II and then to 5 min on stage III caused the slight decrease of MLSS and MLVSS, probably due to the biomass washout in the effluent. Then with the increase of influent pyridine concentration from 2000 mg L−1 to 4000 mg L−1 on stage V, MLSS and MLVSS increased continuously to 5197.4 ± 132.8 mg L−1 and 4608.6 ± 171.1 mg L−1 on day 120, respectively. As the sludge flocs were transformed into the granules, the SVI decreased all the time, from 125.1 ± 12.5 mL g−1 on day 15 to final 25.6 ± 3.6 mL g−1 on day 120 (Fig. 3). However, settling velocity increased all the time, from initial 1.4 ± 0.3 m h−1 to final 37.2 ± 2.7 m h−1 on day 120 (Fig. 3). The settling velocity of the mature granular sludge developed in this study fell in the range of 36.6 ± 8.8 m h−1 reported by Su and Yu [19]. 3.3. Pyridine removal efficiency of pyridine aerobic degrading granules The removal efficiency of pyridine in the SBR system from beginning until the end of granules development period is illustrated in
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3.4. Enhanced pyridine biodegradation by aerobic granules
Fig. 4. Pyridine degradation during 120 days’ operation.
Fig. 4. At stage I of the operation, the pyridine concentration in the effluent decreased from 432.9 ± 30.0 mg L−1 on day 2 to 0 mg L−1 on day 5, presumably due to the adapting process of the biomass with pyridine containing wastewater. 36 days later, with the increase of influent pyridine concentration from 1000 mg L−1 to 1500 mg L−1 and the decrease of the cycle duration from 24 h to 12 h, the pyridine removal efficiency was fluctuating, with the effluent pyridine concentration increased to the maximal 356.6 ± 35.3 mg L−1 on day 36. From then on, pyridine concentration in the effluent decreased, with the effluent quality improved greatly. On day 56, the effluent pyridine concentration increased to 166.7 ± 18.4 mg L−1 suddenly. The deterioration of the reactor performance could be attributed to the wash out of the biomass from the SBR, which was caused by the sudden decrease of the settling time from 5 min to 3 min. Thereafter, as the evolution of flocculent sludge into granular sludge gradually took place in the bioreactor system, the degradation ability for pyridine removal had been improved, with removal efficiency at around 100%, although the influent pyridine concentration was increasing all the time.
The aerobic granules could efficiently degrade pyridine over initial concentration up to 4000 mg L−1 (Fig. 5a). At initial pyridine concentration around 500 mg L−1 , degradation kinetics followed closely zero-order kinetics without obvious lag phase. With the increase of initial pyridine concentration, obvious lag phase was observed, which could be attributed to the high toxicity and recalcitrance of pyridine. However, around 4000 mg L−1 pyridine could be completely degraded within 7.5 h. For the pure Rhizobium sp. NJUST18, although 2600 mg L−1 pyridine could be completely removed in batch reactor, the incubation time needed was as long as 240 h (Fig. 5b). The maximum volumetric degradation rate (Vmax , mg L−1 h−1 ) of pyridine was modeled with the integrated Gompertz equation, according to Shen et al. [9]. The calculated Vmax for pure Rhizobium sp. NJUST18 was in the range from 4.2 mg L h−1 to 32.4 mg L−1 h−1 as the initial pyridine concentrations within the range of 100–2600 mg L−1 . However, in this aerobic granular system, as the initial pyridine concentration within the range of 500–4000 mg L−1 , the calculated Vmax varied between 1164.5 mg L−1 h−1 and 1867.4 mg L−1 h−1 , which was rather high compared with that for pure Rhizobium sp. NJUST18. What is more, there is significant difference in terms of residual TOC concentrations between aerobic granular system and pure Rhizobium sp. NJUST18 system (Fig. 6). As the initial pyridine concentrations within the range of 100–2600 mg L−1 , the residual TOC concentration varied between 65.7 ± 4.2 mg L−1 and 156.8 ± 8.9 mg L−1 in the pure Rhizobium sp. NJUST18 system. As for the aerobic granular system, as the initial pyridine concentrations within the range of 500–4000 mg L−1 , the residual TOC concentration varied between 17.2 ± 4.6 mg L−1 and 74.5 ± 11.8 mg L−1 , which was rather low compared with that for pure Rhizobium sp. NJUST18 system. 3.5. Microbial community structure and dominant species analysis In this study, the microbial community structure of the mature aerobic granules in the SBR system was analyzed by high throughput sequencing based on Illumina Miseq system, with the dominant
Fig. 5. Comparison of pyridine degradation performance between aerobic granules (a) and pure Rhizobium sp. NJUST18 (b).
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Fig. 6. Comparison of TOC removal between aerobic granules and pure Rhizobium sp. NJUST18.
bacteria revealed. As shown in Fig. 7, bacterial sequences affiliated with Proteobacteria (53.41%) were the most abundant, followed by the sequences affiliated with Bacteroidetes (25.12%), unclassified phylum (9.99%), Chlorobium (4.69%), Minor phyla (2.34%), Chloroflexi (2.29%) and Actinobacteria (2.17%). At the genus level, the majority of dominant populations belonged to Paracoccus (55.54%), Comamonas (26.78%), Paludibacter (3.44%), Gemmatimonas (1.65%), Rhizobium (1.21%) and Propionicimonas (1.12%). The bacterial richness and diversity was relatively low, probably due to the selective pressure placed on the bacterial community by high loading of pyridine. 4. Discussion The flagella of NJUST18 cells were assumed to correspond closely to the noted capability to autoaggregate. High autoaggregation of Acinetobacter calcoaceticus was also observed by Adav and Lee [3], the authors attributed the high autoaggregation potential of A. calcoaceticus to the interconnecting fibrils. Strains with flagella, interconnecting fibrils or mycelium were commonly found in the aggregates [20,21]. The inoculation of the aggregating bacterial strains enhanced the granulation process, probably due to the formation of primary matrixes through aggregation [3,5]. Liu and Tay [22] had pointed out that the appearance of primary matrixes was a very crucial step in initiating activated sludge granulation. Sludge granulation could be accelerated if the primary matrix could
be provided rapidly. Thus, it could be inferred that the autoaggregation activity of Rhizobium sp. NJUST18 cells played a significant role in aerobic granulation, probably due to the small and white granules formed within 2 weeks after the startup of this aerobic granular system. In fact, in the pure Rhizobium sp. NJUST18 system, the small and white granules could also be observed, which might served as the primary matrixes during granulation. In addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula was equally important on the aerobic granule formation. The settling property of the mature aerobic granules formed in this study was rather excellent, as was indicated by the low SVI and the high settling velocity values observed after the system had stabilized. The SVI value of 25.6 ± 3.6 mL g−1 at the end of stage VI (day 120) was relatively low, compared with that in the aerobic granule systems described in the literature, which varied in the range of 40–120 mL g−1 [12,23]. This result indicated that the settling properties of the aerobic granules developed in this study were rather good, which was favorable for the operation of the wastewater treatment plant. The high settling velocity of 37.2 m h−1 had given significant impact on the biomass retention in the SBR reactor [18]. Despite the short settling time, the high settling velocity possessed by the developed microbial granular sludge enabled the granules to escape from being flushed out during the decanting phase. Such conditions had caused more microbial granular sludge to be retained in the system and resulted in the increase of biomass concentration in the reactor. The excellent pyridine degradation performance of the mature aerobic granules formed in this study was revealed by both the high Vmax and the high tolerance toward pyridine. The excellent degradation performance in terms of the rather high pyridine removal efficiency indicated the high biological activity occurred during microbial aerobic degradation process of pyridine wastewater in the SBR, even when the influent pyridine concentration was as high as 4000 mg L−1 . The Vmax for mature aerobic granules was much higher than the Vmax for the pure Rhizobium sp. NJUST18. In addition, the residual TOC concentration in the effluent of the aerobic granular system was much lower than that of the pure Rhizobium sp. NJUST18 system. Low residual TOC concentration observed in the aerobic granular system suggested that intermediate metabolite leakage was not evident. It was probably due to the large microbial density, high toxic resistance and relatively high community diversity of aerobic granules. In addition, the compact structure of aerobic granules protected the microbes inside from the inhibitory effects of the target compounds [1]. Aerobic granules reported in this study were able to degrade pyridine at initial concentration up to 4000 mg L−1 within only
Fig. 7. Phylogenetic distribution of sequences assigned on phylum (a) and genus (b).
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7.5 h, which was an advantage for the treatment of pyridine containing wastewater at relatively high concentrations. In a completely mixed activated sludge process seeded with P. pseudoalcaligenes-KPN capable of degrading pyridine, an increase of the influent pyridine concentration to 400 mg L−1 resulted in the reactor failure to degrade pyridine [24]. In an acclimated pyridine biodegradation system based on aerobic granules initially for phenol degradation, at pyridine concentration higher than 3000 mg L−1 , pyridine degradation rate was rather low [8]. The Vmax reported in this literature was as low as 63.7 mg L−1 h−1 , implying a strong inhibitory effect of pyridine on the granules. Although Liu et al. [25] have reported that 4250 mg pyridine L−1 could be degraded completely in an aerobic immobilized microbial system, the hydraulic retention time (HRT) required was as long as 36 h. During the whole operation period of about 6 months in this study, stable treatment performance was observed in the SBR dominated by aerobic granules. This further corroborated the robustness of the granular technology for application in the treatment of wastewaters containing highly toxic and recalcitrant organics. Since the bioaugmented SBR was operated in open condition, other strains from the air or the influent would survive in the bioaugmented SBR. Thus, the significant shift of the microbial community structure was very likely, although the pure Rhizobium sp. NJUST18 was initially inoculated in to the SBR as the sole inocula. It was worth noting that the initially inoculated pyridine degrader, Rhizobium sp. NJUST18, was not the most predominant strain of the pyridine degrading aerobic granules after the system had stabilized. The loss of the dominance of the inoculated strain was consistent with several previous studies. For instance, Wen et al. [26] found that the seeded pyridine degrading strain Paracoccus denitrificans W12 was not the dominant species in a bioaugmented MBR treating pharmaceutical wastewater. Bai et al. [27] also did not find the inoculated bacteria capable of degrading pyridine and quinoline in the bioaugmented zeolite-BAF for coking wastewater treatment. However, in a biaugmented aerobic granular system for 2-fluorphenol degradation, the initially seeded strain capable of degrading 2fluorphenol, i.e., Rhodococcus sp. strain FP1, could be successfully recovered after months of exposure to 2-fluorphenol [28]. Thus, the fate and dynamics of the inoculated degrading bacteria could be the integrated result of various environmental parameters, which would be rather complex [26]. However, considering the rapid and successful start-up of the SBR based aerobic granules, it could be inferred that Rhizobium sp. NJUST18 played a key role in the start-up of this bioaugmentation system. The success of the bioaugmentation by Rhizobium sp. NJUST18 was also confirmed by the complete failure of a control experiment carried out in our lab using the common activated sludge as the sole inoculum for pyridine wastewater treatment. Almost no pyridine was degraded in the control SBR inoculated with activated sludge for the whole experiment period of 2 weeks. In addition, serious biomass flushing was observed in the control SBR, as was indicated by the decreasing MLSS. The complete failure of the control SBR was probably due to the poor biodegradability of pyridine and the high toxicity of pyridine toward the activated sludge. The genera Paracoccus, which was the most predominant in the mature aerobic granules, might be mainly responsible for the pyridine degradation because pyridine was effectively removed in the aerobic granular system at rather high pyridine loading. Previous studies had revealed that Paracoccus sp. was an excellent pyridine degrader [29–31]. The dominance of the genera Comamonas was probably related with its strong aggregation ability. The genus Comamonas had polar flagella and may be contributing to the noted granule stability [32]. Aerobic granulation and phenol degradation in the reactor bioaugmented with Comamonas sp. PG-08 was significantly enhanced compared with a control reactor [5]. Comamonas sp. was also found to be the dominant species of
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the bacterial community in a bioaugmented MBR treating pyridinecontaining pharmaceutical wastewater [26]. In fact, the presence of the aggregation-induced strains could help the formation of microbial aggregates and the removal of recalcitrant compounds, even if they did not participate in critical degradation steps. In our laboratory, six bacterial strains, including Paracoccus thiophilus NJUST24 and Comamonas sp. NJUST25, have been isolated from the mature aerobic granules and used as the inocula for aerobic granulation. Further study is needed to examine the dynamic change of the microbial community structure during the startup and long-term operation of this aerobic granluation system, with the role of Paracoccus sp. and Comamonas sp. emphasized. In addition, other identified bacteria such as Paludibacter and Gemmatimonas, which was also found from a pyridine anoxic degradation system in our laboratory, should also play an important role in pyridine biodegradation. Their detailed function needs to be further studied through selective degradation experiments. 5. Conclusions Cultivation of stable aerobic granules was achieved in a SBR inoculated with an autoaggregative pyridine-degrader, Rhizobium sp. NJUST18. The aerobic granules could degrade pyridine at extremely high Vmax , demonstrating excellent pyridine degradation performance. Bacterial community analysis revealed that Paracoccus and Comamonas were the most predominant species in the aerobic granular system after the system had stabilized. The initially inoculated Rhizobium sp. lost its dominance during aerobic granulation. However, the inoculation of Rhizobium sp. played a key role in the start-up of this bioaugmentation system. In addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula was equally important on the aerobic granule formation. Acknowledgements This research is financed by Major Project of Water Pollution Control and Management Technology of P. R. China(No. 2012ZX07101-003-001), National Natural Science Foundation of China(No. 51478225) and Zijin Intelligent Program of NJUST (No. 2013-ZJ-02-19). References [1] A.M. Maszenan, Y. Liu, W.J. Ng, Bioremediation of wastewaters with recalcitrant organic compounds and metals by aerobic granules, Biotechnol. Adv. 29 (2011) 111–123. [2] L. Liu, G.P. Sheng, W.W. Li, Z.H. Tong, R.J. Zeng, J.X. Liu, J. Xie, S.C. Peng, H.Q. Yu, Cultivation of aerobic granular sludge with a mixed wastewater rich in toxic organics, Biochem. Eng. J. 57 (2011) 7–12. [3] S.S. Adav, D.J. Lee, Single-culture aerobic granules with Acinetobacter calcoaceticus, Appl. Microbiol. Biotechnol. 78 (2008) 551–557. [4] K.L. Ho, B. Lin, Y.Y. Chen, D.J. Lee, Biodegradation of phenol using Corynebacterium sp. DJ1 aerobic granules, Bioresour. Technol. 100 (2009) 5051–5055. [5] H.L. Jiang, J.H. Tay, A.M. Maszenan, S.T.L. Tay, Enhanced phenol biodegradation and aerobic granulation by two coaggregating bacterial strains, Environ. Sci. Technol. 40 (2006) 6137–6142. [6] J.Q. Sun, L. Xu, Y.Q. Tang, F.M. Chen, W.Q. Liu, X.L. Wu, Degradation of pyridine by one Rhodococcus strain in the presence of chromium(VI) or phenol, J. Hazard. Mater. 191 (2011) 62–68. [7] K.V. Padoley, S.N. Mudliar, R.A. Pandey, Heterocyclic nitrogenous pollutants in the environment and their treatment options-an overview, Bioresour. Technol. 99 (2008) 4029–4043. [8] S.S. Adav, D.J. Lee, N.Q. Ren, Biodegradation of pyridine using aerobic granules in the presence of phenol, Water Res. 41 (2007) 2903–2910. [9] J.Y. Shen, X. Zhang, D. Chen, X.D. Liu, L.B. Zhang, X.Y. Sun, J.S. Li, H.P. Bi, L.J. Wang, Kinetics study of pyridine biodegradation by a novel bacterial strain, Rhizobium sp. NJUST18, Bioprocess Biosyst. Eng. 37 (2014) 1185–1192. [10] J.Y. Shen, J.F. Zhang, Y. Zuo, L.J. Wang, X.Y. Sun, J.S. Li, W.Q. Han, R. He, Biodegradation of 2 4,6-trinitrophenol by Rhodococcus sp. isolated from a picric acid-contaminated soil, J. Hazard. Mater. 163 (2009) 1199–1206.
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[11] APHA, Standard Methods for the Examination of Water and Wastewater, 19th edition, American Public Health Association, Washington DC, 1998. [12] K. Muda, A. Aris, M.R. Salim, Z. Ibrahim, A. Yahya, M.C.M. van Loosdrecht, A. Ahmad, Z. Nawahwi, Development of granular sludge for textile wastewater treatment, Water Res. 44 (2010) 4341–4350. [13] Y.J. Song, S. Ishii, L. Rathnayake, T. Ito, H. Satoh, S. Okabe, Development and characterization of the partial nitrification aerobic granules in a sequencing batch airlift reactor, Bioresour. Technol. 139 (2013) 285–291. [14] A. Bhattacharya, A. Gupta, A. Kaur, D. Malik, Efficacy of Acinetobacter sp. B9 for simultaneous removal of phenol and hexavalent chromium from co-contaminated system, Appl. Microbiol. Biotechnol. (2014), http://dx.doi.org/10.1007/s00253-014-5910-5. [15] B. Liang, H.Y. Cheng, J.D.V. Nostrand, J.C. Ma, H. Yu, D.Y. Kong, W.Z. Liu, N.Q. Ren, L.Y. Wu, A.J. Wang, D.J. Lee, J.Z. Zhou, Microbial community structure and function of nitrobenzene reduction biocathode in response to carbon source switchover, Water Res. 54 (2014) 137–148. [16] I. Pala-Ozkok, A. Rehman, N. Yagci, E. Ubay-Cokgor, D. Jonas, D. Orhon, Characteristics of mixed microbial culture at different sludge ages: effect on variable kinetics for substrate utilization, Bioresour. Technol. 126 (2012) 274–282. [17] E. Dulekgurgen, N. Artan, D. Orhon, P.A. Wilderer, How does shear affect aggregation in granular sludge sequencing batch reactors? Relations between shear hydrophobicity and extracellular polymeric substances, Water Sci. Technol. 58 (2008) 267–276. [18] N.H. Rosman, A.N. Anuar, I. Othman, H. Harun, M.Z. Sulong, S.H. Elias, M.A.H.M. Hassan, S. Chelliapan, Z. Ujang, Cultivation of aerobic granular sludge for rubber wastewater treatment, Bioresour. Technol. 129 (2013) 620–623. [19] K.Z. Su, H.Q. Yu, Formation and characterization of aerobic granules in a sequencing batch reactor treating soybean-processing wastewater, Environ. Sci. Technol. 39 (2005) 2818–2827. [20] S.S. Adav, D.J. Lee, K.Y. Show, J.H. Tay, Aerobic granular sludge: recent advances, Biotechnol. Adv. 26 (2008) 411–423. [21] H.L. Wang, L. Li, P. Li, H. Li, G.S. Liu, J.M. Yao, The acceleration of sludge granulation using the chlamydospores of Phanerochaete sp. HSD, J. Hazard. Mater. 192 (2011) 963–969.
[22] Y. Liu, J.H. Tay, State of the art of biogranulation technology for wastewater treatment, Biotechnol. Adv. 22 (2004) 533–563. [23] B.S. McSwain, R.L. Irvine, P.A. Wilderer, The influence of settling time on the formation of aerobic granules, Water Sci. Technol. 50 (2004) 195–202. [24] K.V. Padoley, A.S. Rajvaidya, T.V. Subbarao, R.A. Pandey, Biodegradation of pyridine in a completely mixed activated sludge process, Bioresour. Technol. 97 (2006) 1225–1236. [25] G. Liu, Z. Ye, H. Li, R. Che, L. Cui, Biological treatment of hexanitrostilbene (HNS) produced wastewater using an anaerobic–aerobic immobilized microbial system, Chem. Eng. J. 213 (2012) 118–124. [26] D.H. Wen, J. Zhang, R.L. Xiong, R. Liu, L.J. Chen, Bioaugmentation with a pyridine-degrading bacterium in a membrane bioreactor treating pharmaceutical wastewater, J. Environ. Sci. 25 (2013) 2265–2271. [27] Y.H. Bai, Q.H. Sun, R.H. Sun, D.H. Wen, X.Y. Tang, Bioaugmentation and adsorption treatment of coking wastewater containing pyridine and quinoline using zeolite-biological aerated filters, Environ. Sci. Technol. 45 (2011) 1940–1948. [28] A.F. Duque, V.S. Bessa, M.F. Carvalho, M.K. de Kreuk, M.C.M. van Loosdrecht, P.M.L. Castro, 2-Fluorophenol degradation by aerobic granular sludge in a sequencing batch reactor, Water Res. 45 (2011) 6745–6752. [29] Y.H. Bai, Q.H. Sun, C. Zhao, D.H. Wen, X.Y. Tang, Microbial degradation and metabolic pathway of pyridine by a Paracoccus sp. strain BW001, Biodegradation 19 (2008) 915–926. [30] L. Qiao, D.H. Wen, J.L. Wang, Biodegradation of pyridine by Paracoccus sp. KT-5 immobilized on bamboo-based activated carbon, Bioresour. Technol. 101 (2010) 5229–5234. [31] L. Qiao, J.L. Wang, Microbial degradation of pyridine by Paracoccus sp. isolated from contaminated soil, J. Hazard. Mater. 176 (2010) 220–225. [32] Y. Lv, C. Wan, X. Liu, Y. Zhang, D.J. Lee, J.H. Tay, Freezing of aerobic granules for storage and subsequent recovery, J. Taiwan Inst. Chem. Eng. 44 (2013) 770–773.
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Aerobic granulation strategy for bioaugmentation of a sequencing batch reactor (SBR) treating high strength pyridine wastewater Xiaodong Liu a , Yan Chen a , Xin Zhang a,b , Xinbai Jiang a , Shijing Wu a , Jinyou Shen a,∗ , Xiuyun Sun a , Jiansheng Li a , Lude Lu a , Lianjun Wang a,∗∗ a Jiangsu Key Laboratory for Chemical Pollution Control and Resources Reuse, School of Environmental and Biological Engineering, Nanjing University of Science and Technology, Nanjing 210094, Jiangsu Province, China b Suzhou Institute of Architectural Design Co., Ltd, Suzhou 215021, Jiangsu Province, China
g r a p h i c a l
a r t i c l e
a b s t r a c t
i n f o
Article history: Received 26 August 2014 Received in revised form 6 March 2015 Accepted 9 April 2015 Available online 13 April 2015 Keywords: Aerobic granule Bioaugmentation Biodegradation High-throughput sequencing
a b s t r a c t Aerobic granules were successfully cultivated in a sequencing batch reactor (SBR), using a single bacterial strain Rhizobium sp. NJUST18 as the inoculum. NJUST18 presented as both a good pyridine degrader and an efficient autoaggregator. Stable granules with diameter of 0.5–1 mm, sludge volume index of 25.6 ± 3.6 mL g−1 and settling velocity of 37.2 ± 2.7 m h−1 , were formed in SBR following 120-day cultivation. These granules exhibited excellent pyridine degradation performance, with maximum volumetric degradation rate (Vmax ) varied between 1164.5 mg L−1 h−1 and 1867.4 mg L−1 h−1 . High-throughput sequencing analysis exhibited a large shift in microbial community structure, since the SBR was operated under open condition. Paracoccus and Comamonas were found to be the most predominant species in the aerobic granule system after the system had stabilized. The initially inoculated Rhizobium sp. lost its dominance during aerobic granulation. However, the inoculation of Rhizobium sp. played a key role in the startup process of this bioaugmentation system. This study demonstrated that, in addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula is equally important for the formation of the aerobic granule. © 2015 Elsevier B.V. All rights reserved.
∗ Corresponding author. Tel. +86 25 84303965; fax: +86 25 84303965. ∗∗ Corresponding author. Tel. +86 25 84315941; fax: +86 25 84315941. E-mail addresses: [email protected] (J. Shen), [email protected] (L. Wang). http://dx.doi.org/10.1016/j.jhazmat.2015.04.025 0304-3894/© 2015 Elsevier B.V. All rights reserved.
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1. Introduction Aerobic granules have been employed in treating toxic and recalcitrant organics, such as phenol, p-nitrophenol and 2,4dichlorophenol, exhibiting high resistance to these toxic and recalcitrant compounds [1]. The high toxic resistance and degradation ability of the aerobic granules are partially attributed to the unique structure of granules, where the dense microbial cells and extracellular polymeric substances (EPS) set a barrier for mass transfer and lower the concentration of toxics on the inner cells [2]. Considering the compact structure, large microbial density and high toxic resistance, aerobic granules might be the core of bioaugmentation [1]. In the formation of aerobic granules, hydraulic selection pressure to select and retain compact microbial aggregates plays a key role. In addition, cell aggregation including autoaggregative or coaggregative interactions may be promoted for aerobic granulation. In the previous studies, aerobic granulation from coaggregative or autoaggregative phenol-degrading strains has been reported for enhanced phenol biodegradation [3–5]. However, detailed information regarding the roles that coaggregation/autoaggregation played in aerobic granulation has not yet been fully investigated. Although, several works have been reported to isolate efficient microbial species for pyridine biodegradation [6], most of the work reported on this heterocyclic pollutant degradation pertained to flask level experiments. Only limited reports with technological orientation are available by now [7]. Reports on aerobic granulation aiming at the bioremediation of pyridine waste are rare [8]. In addition, formation of pyridine-degrading aerobic granules from pure cultures capable of degrading pyridine has not previously been explored to our knowledge. Therefore, the present study aimed to investigate the feasibility of aerobic granulation from Rhizobium sp. NJUST18, which was both a functional pyridine degrader and a good autoaggregator. The effectiveness of the aerobic granules in treating high strength pyridine wastewater was evaluated, and the bacterial community structure of the aerobic granules was also included. 2. Materials and methods
bioreactor was operated in a thermostatic bath to keep cultivation temperature stable at 30 ◦ C. The SBR was initially seeded with 2 g (measured in dry weight) Rhizobium sp. NJUST18 pure culture and fed with the synthetic influent. After inoculation, the SBR was operated in open condition, with the influent and the air introduced without sterilization. Operational parameters for different phases of the cultivation procedure are presented in Table 1. During the start-up period, i.e., stage I (1–35 days), the bioreactor was operated at a cycle of 24 h:5 min of feeding without stirring, 23.5 h of aerobic reaction, 12 min of settling, 3 min of effluent withdrawal and 10 min of idling. For the first week, in addition to 1000 mg L−1 pyridine, 1000 mg L−1 acetate sodium was added into synthetic influent as supplementary substrate for biomass growth stimulation. One week later, acetate sodium in synthetic influent was reduced progressively, and finally reduced to 0 mg L−1 on day 30. Thereafter, influent pyridine concentration increased to 4000 mg L−1 step by step, while the cycle time and settling time progressively decreased to 6 h and 2 min, respectively. 120 days later, mature granules were observed in SBR, with stable operation of the SBR achieved. In order to investigate the effect of pyridine concentration on pyridine biodegradation in the bioaugmented SBR, 1000–8000 mg L−1 pyridine was added into the synthetic influent to make initial pyridine in SBR over the concentration range between 500 and 4000 mg L−1 . In order to reveal the key role of Rhizobium sp. NJUST18 during the startup of the bioaugmented SBR, a SBR inoculated with common activated sludge was operated as a control. The common activated sludge used was taken from the secondary sedimentation tank of a municipal sewage treatment plant located in Nanjing. The control SBR was initially seeded with 6 g activated sludge (measured in dry weight) and was operated under the same conditions as the bioaugmented SBR, except for the inocula used. Specifically, 1000 mg L−1 pyridine and 1000 mg L−1 acetate sodium was added into the synthetic influent as the carbon sources for the first week. One week later, acetate sodium in synthetic influent was reduced to 800 mg L−1 while pyridine remained the same as the first week. However, due to the complete failure, the control SBR stopped running after 2 weeks.
2.1. Microorganism and synthetic wastewater The isolation and cultivation of Rhizobium sp. NJUST18 was described in our previous study [9]. The 16S rRNA sequence (comprising 1381 nucleotides) was deposited in the GenBank database under accession no. JN106368. The synthetic influent of the sequential batch reactor (SBR) was as follows: Na2 HPO4 • 12H2 O (3.057 g L−1 ), KH2 PO4 (0.743 g L−1 ), NH4 Cl (1 g L−1 ), MgSO4 • 7H2 O (0.89 g L−1 ), KCl (0.35 g L−1 ), CaCl2 (0.20 g L−1 ), FeCl3 (0.03 g L−1 ), pyridine and sodium acetate at desired concentrations, and 10 mL L−1 trace element solution SL-4 [10]. 2.2. Aerobic granules cultivation The aerobic granules were cultivated in a column-type sequential batch reactor (SBR) with inner diameter of 6 cm, height of 100 cm and effective volume of 2.2 L. The sequential operation of the reactors was automatically controlled by the timers. During the filling phase, the synthetic influent was introduced through ports located at the bottom of the bioreactor. While, fine air bubbles for aeration were supplied by means of air bubble diffusers placed at the bottom of the SBR at volumetric flow rate of 0.2 m3 h−1 and superficial gas velocity of 0.02 m s−1 . The effluent was withdrawn through the outlet ports positioned at the medium height in this column reactor, resulting in volumetric exchange ratio of 50%. The
2.3. Analytical methods Pyridine was quantified by HPLC (Waters 2996, Waters Incorporation, USA) through authentic standard. Before analysis, water samples were passed through a 0.22 m filter. The HPLC analysis was conducted at room temperature using a Waters RP18 column (5 m, 4.6 × 250 mm) and a UV–vis detector. The mobile phase was a mixture of 70% methanol and 30% water pumped at a flow rate of 1.00 mL/min. The analysis was performed at 254 nm, with column temperature at 35 ◦ C. Measurements of the parameters, such as sludge volume index (SVI), mixed liquor suspended solid (MLSS) and mixed liquor volatile suspended solids (MLVSS), were carried out according to APHA standard methods [11]. The settling velocity of the sludge was measured by recording the traveling time of individual granules from the top to the bottom of a cylinder [12]. The granule size was manually measured using a fine-scale ruler [13]. Scanning electron microscopy (SEM) observation was carried out according to Ho et al. [4]. Transmission electron microscopy (TEM) observation was carried out according to Bhattacharya et al. [14]. The effluent samples for the determination of the residual TOC concentration were sampled 2 h after the complete exhaustion of the pyridine. TOC concentration was measured using a Germany Elementar vario TOC analyzer.
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Table 1 Operational parameters applied during different cultivation phases. Stage
I
II
III
IV
V
VI
Time (day) Influent pyridine concentration (mg L−1 ) Influent NaAC concentration (mg L−1 ) Cycle duration (h) Settling time (min) Aeration time (min) Pyridine loading rate (kg m−3 d−1 )
1–35 1000 1000–0 24 12 1410 0.5
36–45 1500 0 12 10 692 1.5
46–55 2000 0 12 5 697 2.0
56–65 2000 0 8 3 459 3.0
66–85 4000 0 6 3 339 8.0
86–120 4000 0 6 2 340 8.0
2.4. DNA extraction, PCR and High-throughput sequencing The microbial community structure of the aerobic granules in the SBR system was profiled using high-throughput sequencing technology. About 1 g sample of mature aerobic granules was
collected at the end of stage VI. DNA of the aerobic granules was extracted from the aerobic granule sample using the FastDNA® SPIN for soil kit (MP Biomedicals, CA, USA) according to the manufacturer’s instruction. The concentration and purity of genomic DNA were measured using NanoDrop® ND-1000 (NanoDrop Technolo-
Fig. 1. TEM of Rhizobium sp. NJUST18 (a), biomass 2 weeks after inoculation (b), biomass 14 weeks after inoculation (c), biomass18 weeks after inoculation (d), and SEM of mature granule (e and f).
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Fig. 2. The MLSS, MLVSS and settling time profiles during 120 days’ operation.
gies, Willmington, DE, USA) to ensure the sample concentration higher than 20ng/L and A260/A280 between 1.80 and 2.00. The extracted DNA was amplified by polymerase chain reaction (PCR) using TaKaRa Ex Taq® (TaKaRa Bio Japan) according to the manufacturer’s instruction. Primers 27F (5 -AGAGTTTGATYMTGGCTCAG-3 ) and 338R (5 TGCTGCCTCCCGTAGGAGT-3 ) which targeted V1/V2 hypervariable regions of bacterial 16S rRNA genes were selected. In order to tag the PCR products from each sample, specific 8 bases long identifier were included in both forward and reverse primers. For sequencing the PCR products, specific sequencing adaptors were attached to both primers. The extracted DNA was diluted to 20ng/L and used as the template DNA. Thermal cycling conditions were as follows: an initial denaturation at 98 ◦ C for 5 min, and 20 cycles at 98 ◦ C for 30 s, 50 ◦ C for 30 s, and 72 ◦ C for 40 s, with a final extension at 72 ◦ C for 10 min. Four parallel PCR products were pooled and purified by OMEGA E.Z.N.A Cycle-Pure Kit (Omega Bio-Tek Norcross, USA). The purity of purified PCR product was measured using NanoDrop® ND-1000 once again to ensure its A260/A280 between 1.70 and 1.90. The purified PCR product was quantified precisely using Qubit 2.0 Fluorometer (Life Technologies, Grand Island, USA). The purified library was diluted, denatured, rediluted, mixed with PhiX (equal to 30% of final DNA amount) as described in the Illumina library preparation protocols, and then applied to an Illumina Miseq system in Jiangsu Zhongyijinda Analytical & Testing Co., Ltd. for sequencing with the Reagent Kit v22 × 250 bp as described in the manufacture manual. DNA library building and data analysis were performed in according to Liang et al. [15] and Pala-Ozkok et al. [16]. 3. Results 3.1. Formation of pyridine aerobic degrading granule TEM image of Rhizobium sp. NJUST18 (Fig. 1a) showed that strain NJUST18 was a rod-shaped bacterium with flagella, which were assumed to correspond closely to the noted capability of autoaggregation. One week after the inoculation of the autoaggregative NJUST18, both biomass increase and obvious aggregation in SBR were observed. Two weeks later, the biomass in SBR showed fluffy, irregular, loose-structure morphology (Fig. 1b). Some small and white granules could be observed, however, these granules were easily broke up into small pieces if placed under vigorous shaking. 14 weeks after inoculation, with the increase of influent pyridine concentration, decrease of the cycle time and settling time, the flocs-like sludge gradually decreased and was replaced by small
Fig. 3. The settling velocities and SVI profiles during 120 days’ operation.
granules with diameter of 0.2–0.5 mm, while the color of the granules gradually changed from white to yellowish brown (Fig. 1c). In the following 4 weeks, the granules became denser and more regular in shape under high shear force which induced the biomass aggregates to secrete more exopolysaccharides [17,18]. Through the control of settling time, more flocculent sludge was washed out from the SBR, resulting in the accumulation of the aerobic granules with high settling velocity. After 18 weeks of inoculation, mature granules (Fig. 1d) with size of 0.5–1 mm were formed, leading to a stable operation of the SBR. The mature granules turned out to be smooth, with a solid surface (Fig. 1d and e). Through SEM observation, rod-shaped bacteria were found to be dominated at the granule surface (Fig. 1f), indicating that rod-shaped strains played an important role in the granule formation and granule stabilization. 3.2. Biomass profile and settling properties of pyridine aerobic degrading granules Since the settling time is a critical parameter and significantly influences the granulation process, a short settling time was generally considered favorable for the granule formation [2]. Fig. 2 shows the MLSS, MLVSS during the development period of the pyridine aerobic degrading granules. During the stage I, the settling time was fixed at 12 min. MLSS and MLVSS increased from initial 878.2 ± 30.8 mg L−1 and 853.6 ± 45.3 mg L−1 to 3582.0 ± 63.8 mg L−1 and 3451.1 ± 194.5 mg L−1 on day 30, respectively. Decreasing the settling time from 12 min to 10 min on stage II and then to 5 min on stage III caused the slight decrease of MLSS and MLVSS, probably due to the biomass washout in the effluent. Then with the increase of influent pyridine concentration from 2000 mg L−1 to 4000 mg L−1 on stage V, MLSS and MLVSS increased continuously to 5197.4 ± 132.8 mg L−1 and 4608.6 ± 171.1 mg L−1 on day 120, respectively. As the sludge flocs were transformed into the granules, the SVI decreased all the time, from 125.1 ± 12.5 mL g−1 on day 15 to final 25.6 ± 3.6 mL g−1 on day 120 (Fig. 3). However, settling velocity increased all the time, from initial 1.4 ± 0.3 m h−1 to final 37.2 ± 2.7 m h−1 on day 120 (Fig. 3). The settling velocity of the mature granular sludge developed in this study fell in the range of 36.6 ± 8.8 m h−1 reported by Su and Yu [19]. 3.3. Pyridine removal efficiency of pyridine aerobic degrading granules The removal efficiency of pyridine in the SBR system from beginning until the end of granules development period is illustrated in
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3.4. Enhanced pyridine biodegradation by aerobic granules
Fig. 4. Pyridine degradation during 120 days’ operation.
Fig. 4. At stage I of the operation, the pyridine concentration in the effluent decreased from 432.9 ± 30.0 mg L−1 on day 2 to 0 mg L−1 on day 5, presumably due to the adapting process of the biomass with pyridine containing wastewater. 36 days later, with the increase of influent pyridine concentration from 1000 mg L−1 to 1500 mg L−1 and the decrease of the cycle duration from 24 h to 12 h, the pyridine removal efficiency was fluctuating, with the effluent pyridine concentration increased to the maximal 356.6 ± 35.3 mg L−1 on day 36. From then on, pyridine concentration in the effluent decreased, with the effluent quality improved greatly. On day 56, the effluent pyridine concentration increased to 166.7 ± 18.4 mg L−1 suddenly. The deterioration of the reactor performance could be attributed to the wash out of the biomass from the SBR, which was caused by the sudden decrease of the settling time from 5 min to 3 min. Thereafter, as the evolution of flocculent sludge into granular sludge gradually took place in the bioreactor system, the degradation ability for pyridine removal had been improved, with removal efficiency at around 100%, although the influent pyridine concentration was increasing all the time.
The aerobic granules could efficiently degrade pyridine over initial concentration up to 4000 mg L−1 (Fig. 5a). At initial pyridine concentration around 500 mg L−1 , degradation kinetics followed closely zero-order kinetics without obvious lag phase. With the increase of initial pyridine concentration, obvious lag phase was observed, which could be attributed to the high toxicity and recalcitrance of pyridine. However, around 4000 mg L−1 pyridine could be completely degraded within 7.5 h. For the pure Rhizobium sp. NJUST18, although 2600 mg L−1 pyridine could be completely removed in batch reactor, the incubation time needed was as long as 240 h (Fig. 5b). The maximum volumetric degradation rate (Vmax , mg L−1 h−1 ) of pyridine was modeled with the integrated Gompertz equation, according to Shen et al. [9]. The calculated Vmax for pure Rhizobium sp. NJUST18 was in the range from 4.2 mg L h−1 to 32.4 mg L−1 h−1 as the initial pyridine concentrations within the range of 100–2600 mg L−1 . However, in this aerobic granular system, as the initial pyridine concentration within the range of 500–4000 mg L−1 , the calculated Vmax varied between 1164.5 mg L−1 h−1 and 1867.4 mg L−1 h−1 , which was rather high compared with that for pure Rhizobium sp. NJUST18. What is more, there is significant difference in terms of residual TOC concentrations between aerobic granular system and pure Rhizobium sp. NJUST18 system (Fig. 6). As the initial pyridine concentrations within the range of 100–2600 mg L−1 , the residual TOC concentration varied between 65.7 ± 4.2 mg L−1 and 156.8 ± 8.9 mg L−1 in the pure Rhizobium sp. NJUST18 system. As for the aerobic granular system, as the initial pyridine concentrations within the range of 500–4000 mg L−1 , the residual TOC concentration varied between 17.2 ± 4.6 mg L−1 and 74.5 ± 11.8 mg L−1 , which was rather low compared with that for pure Rhizobium sp. NJUST18 system. 3.5. Microbial community structure and dominant species analysis In this study, the microbial community structure of the mature aerobic granules in the SBR system was analyzed by high throughput sequencing based on Illumina Miseq system, with the dominant
Fig. 5. Comparison of pyridine degradation performance between aerobic granules (a) and pure Rhizobium sp. NJUST18 (b).
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Fig. 6. Comparison of TOC removal between aerobic granules and pure Rhizobium sp. NJUST18.
bacteria revealed. As shown in Fig. 7, bacterial sequences affiliated with Proteobacteria (53.41%) were the most abundant, followed by the sequences affiliated with Bacteroidetes (25.12%), unclassified phylum (9.99%), Chlorobium (4.69%), Minor phyla (2.34%), Chloroflexi (2.29%) and Actinobacteria (2.17%). At the genus level, the majority of dominant populations belonged to Paracoccus (55.54%), Comamonas (26.78%), Paludibacter (3.44%), Gemmatimonas (1.65%), Rhizobium (1.21%) and Propionicimonas (1.12%). The bacterial richness and diversity was relatively low, probably due to the selective pressure placed on the bacterial community by high loading of pyridine. 4. Discussion The flagella of NJUST18 cells were assumed to correspond closely to the noted capability to autoaggregate. High autoaggregation of Acinetobacter calcoaceticus was also observed by Adav and Lee [3], the authors attributed the high autoaggregation potential of A. calcoaceticus to the interconnecting fibrils. Strains with flagella, interconnecting fibrils or mycelium were commonly found in the aggregates [20,21]. The inoculation of the aggregating bacterial strains enhanced the granulation process, probably due to the formation of primary matrixes through aggregation [3,5]. Liu and Tay [22] had pointed out that the appearance of primary matrixes was a very crucial step in initiating activated sludge granulation. Sludge granulation could be accelerated if the primary matrix could
be provided rapidly. Thus, it could be inferred that the autoaggregation activity of Rhizobium sp. NJUST18 cells played a significant role in aerobic granulation, probably due to the small and white granules formed within 2 weeks after the startup of this aerobic granular system. In fact, in the pure Rhizobium sp. NJUST18 system, the small and white granules could also be observed, which might served as the primary matrixes during granulation. In addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula was equally important on the aerobic granule formation. The settling property of the mature aerobic granules formed in this study was rather excellent, as was indicated by the low SVI and the high settling velocity values observed after the system had stabilized. The SVI value of 25.6 ± 3.6 mL g−1 at the end of stage VI (day 120) was relatively low, compared with that in the aerobic granule systems described in the literature, which varied in the range of 40–120 mL g−1 [12,23]. This result indicated that the settling properties of the aerobic granules developed in this study were rather good, which was favorable for the operation of the wastewater treatment plant. The high settling velocity of 37.2 m h−1 had given significant impact on the biomass retention in the SBR reactor [18]. Despite the short settling time, the high settling velocity possessed by the developed microbial granular sludge enabled the granules to escape from being flushed out during the decanting phase. Such conditions had caused more microbial granular sludge to be retained in the system and resulted in the increase of biomass concentration in the reactor. The excellent pyridine degradation performance of the mature aerobic granules formed in this study was revealed by both the high Vmax and the high tolerance toward pyridine. The excellent degradation performance in terms of the rather high pyridine removal efficiency indicated the high biological activity occurred during microbial aerobic degradation process of pyridine wastewater in the SBR, even when the influent pyridine concentration was as high as 4000 mg L−1 . The Vmax for mature aerobic granules was much higher than the Vmax for the pure Rhizobium sp. NJUST18. In addition, the residual TOC concentration in the effluent of the aerobic granular system was much lower than that of the pure Rhizobium sp. NJUST18 system. Low residual TOC concentration observed in the aerobic granular system suggested that intermediate metabolite leakage was not evident. It was probably due to the large microbial density, high toxic resistance and relatively high community diversity of aerobic granules. In addition, the compact structure of aerobic granules protected the microbes inside from the inhibitory effects of the target compounds [1]. Aerobic granules reported in this study were able to degrade pyridine at initial concentration up to 4000 mg L−1 within only
Fig. 7. Phylogenetic distribution of sequences assigned on phylum (a) and genus (b).
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7.5 h, which was an advantage for the treatment of pyridine containing wastewater at relatively high concentrations. In a completely mixed activated sludge process seeded with P. pseudoalcaligenes-KPN capable of degrading pyridine, an increase of the influent pyridine concentration to 400 mg L−1 resulted in the reactor failure to degrade pyridine [24]. In an acclimated pyridine biodegradation system based on aerobic granules initially for phenol degradation, at pyridine concentration higher than 3000 mg L−1 , pyridine degradation rate was rather low [8]. The Vmax reported in this literature was as low as 63.7 mg L−1 h−1 , implying a strong inhibitory effect of pyridine on the granules. Although Liu et al. [25] have reported that 4250 mg pyridine L−1 could be degraded completely in an aerobic immobilized microbial system, the hydraulic retention time (HRT) required was as long as 36 h. During the whole operation period of about 6 months in this study, stable treatment performance was observed in the SBR dominated by aerobic granules. This further corroborated the robustness of the granular technology for application in the treatment of wastewaters containing highly toxic and recalcitrant organics. Since the bioaugmented SBR was operated in open condition, other strains from the air or the influent would survive in the bioaugmented SBR. Thus, the significant shift of the microbial community structure was very likely, although the pure Rhizobium sp. NJUST18 was initially inoculated in to the SBR as the sole inocula. It was worth noting that the initially inoculated pyridine degrader, Rhizobium sp. NJUST18, was not the most predominant strain of the pyridine degrading aerobic granules after the system had stabilized. The loss of the dominance of the inoculated strain was consistent with several previous studies. For instance, Wen et al. [26] found that the seeded pyridine degrading strain Paracoccus denitrificans W12 was not the dominant species in a bioaugmented MBR treating pharmaceutical wastewater. Bai et al. [27] also did not find the inoculated bacteria capable of degrading pyridine and quinoline in the bioaugmented zeolite-BAF for coking wastewater treatment. However, in a biaugmented aerobic granular system for 2-fluorphenol degradation, the initially seeded strain capable of degrading 2fluorphenol, i.e., Rhodococcus sp. strain FP1, could be successfully recovered after months of exposure to 2-fluorphenol [28]. Thus, the fate and dynamics of the inoculated degrading bacteria could be the integrated result of various environmental parameters, which would be rather complex [26]. However, considering the rapid and successful start-up of the SBR based aerobic granules, it could be inferred that Rhizobium sp. NJUST18 played a key role in the start-up of this bioaugmentation system. The success of the bioaugmentation by Rhizobium sp. NJUST18 was also confirmed by the complete failure of a control experiment carried out in our lab using the common activated sludge as the sole inoculum for pyridine wastewater treatment. Almost no pyridine was degraded in the control SBR inoculated with activated sludge for the whole experiment period of 2 weeks. In addition, serious biomass flushing was observed in the control SBR, as was indicated by the decreasing MLSS. The complete failure of the control SBR was probably due to the poor biodegradability of pyridine and the high toxicity of pyridine toward the activated sludge. The genera Paracoccus, which was the most predominant in the mature aerobic granules, might be mainly responsible for the pyridine degradation because pyridine was effectively removed in the aerobic granular system at rather high pyridine loading. Previous studies had revealed that Paracoccus sp. was an excellent pyridine degrader [29–31]. The dominance of the genera Comamonas was probably related with its strong aggregation ability. The genus Comamonas had polar flagella and may be contributing to the noted granule stability [32]. Aerobic granulation and phenol degradation in the reactor bioaugmented with Comamonas sp. PG-08 was significantly enhanced compared with a control reactor [5]. Comamonas sp. was also found to be the dominant species of
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the bacterial community in a bioaugmented MBR treating pyridinecontaining pharmaceutical wastewater [26]. In fact, the presence of the aggregation-induced strains could help the formation of microbial aggregates and the removal of recalcitrant compounds, even if they did not participate in critical degradation steps. In our laboratory, six bacterial strains, including Paracoccus thiophilus NJUST24 and Comamonas sp. NJUST25, have been isolated from the mature aerobic granules and used as the inocula for aerobic granulation. Further study is needed to examine the dynamic change of the microbial community structure during the startup and long-term operation of this aerobic granluation system, with the role of Paracoccus sp. and Comamonas sp. emphasized. In addition, other identified bacteria such as Paludibacter and Gemmatimonas, which was also found from a pyridine anoxic degradation system in our laboratory, should also play an important role in pyridine biodegradation. Their detailed function needs to be further studied through selective degradation experiments. 5. Conclusions Cultivation of stable aerobic granules was achieved in a SBR inoculated with an autoaggregative pyridine-degrader, Rhizobium sp. NJUST18. The aerobic granules could degrade pyridine at extremely high Vmax , demonstrating excellent pyridine degradation performance. Bacterial community analysis revealed that Paracoccus and Comamonas were the most predominant species in the aerobic granular system after the system had stabilized. The initially inoculated Rhizobium sp. lost its dominance during aerobic granulation. However, the inoculation of Rhizobium sp. played a key role in the start-up of this bioaugmentation system. In addition to the hydraulic selection pressure during settling and effluent discharge, the selection of aggregating bacterial inocula was equally important on the aerobic granule formation. Acknowledgements This research is financed by Major Project of Water Pollution Control and Management Technology of P. R. China(No. 2012ZX07101-003-001), National Natural Science Foundation of China(No. 51478225) and Zijin Intelligent Program of NJUST (No. 2013-ZJ-02-19). References [1] A.M. Maszenan, Y. Liu, W.J. Ng, Bioremediation of wastewaters with recalcitrant organic compounds and metals by aerobic granules, Biotechnol. Adv. 29 (2011) 111–123. [2] L. Liu, G.P. Sheng, W.W. Li, Z.H. Tong, R.J. Zeng, J.X. Liu, J. Xie, S.C. Peng, H.Q. Yu, Cultivation of aerobic granular sludge with a mixed wastewater rich in toxic organics, Biochem. Eng. J. 57 (2011) 7–12. [3] S.S. Adav, D.J. Lee, Single-culture aerobic granules with Acinetobacter calcoaceticus, Appl. Microbiol. Biotechnol. 78 (2008) 551–557. [4] K.L. Ho, B. Lin, Y.Y. Chen, D.J. Lee, Biodegradation of phenol using Corynebacterium sp. DJ1 aerobic granules, Bioresour. Technol. 100 (2009) 5051–5055. [5] H.L. Jiang, J.H. Tay, A.M. Maszenan, S.T.L. Tay, Enhanced phenol biodegradation and aerobic granulation by two coaggregating bacterial strains, Environ. Sci. Technol. 40 (2006) 6137–6142. [6] J.Q. Sun, L. Xu, Y.Q. Tang, F.M. Chen, W.Q. Liu, X.L. Wu, Degradation of pyridine by one Rhodococcus strain in the presence of chromium(VI) or phenol, J. Hazard. Mater. 191 (2011) 62–68. [7] K.V. Padoley, S.N. Mudliar, R.A. Pandey, Heterocyclic nitrogenous pollutants in the environment and their treatment options-an overview, Bioresour. Technol. 99 (2008) 4029–4043. [8] S.S. Adav, D.J. Lee, N.Q. Ren, Biodegradation of pyridine using aerobic granules in the presence of phenol, Water Res. 41 (2007) 2903–2910. [9] J.Y. Shen, X. Zhang, D. Chen, X.D. Liu, L.B. Zhang, X.Y. Sun, J.S. Li, H.P. Bi, L.J. Wang, Kinetics study of pyridine biodegradation by a novel bacterial strain, Rhizobium sp. NJUST18, Bioprocess Biosyst. Eng. 37 (2014) 1185–1192. [10] J.Y. Shen, J.F. Zhang, Y. Zuo, L.J. Wang, X.Y. Sun, J.S. Li, W.Q. Han, R. He, Biodegradation of 2 4,6-trinitrophenol by Rhodococcus sp. isolated from a picric acid-contaminated soil, J. Hazard. Mater. 163 (2009) 1199–1206.
160
X. Liu et al. / Journal of Hazardous Materials 295 (2015) 153–160
[11] APHA, Standard Methods for the Examination of Water and Wastewater, 19th edition, American Public Health Association, Washington DC, 1998. [12] K. Muda, A. Aris, M.R. Salim, Z. Ibrahim, A. Yahya, M.C.M. van Loosdrecht, A. Ahmad, Z. Nawahwi, Development of granular sludge for textile wastewater treatment, Water Res. 44 (2010) 4341–4350. [13] Y.J. Song, S. Ishii, L. Rathnayake, T. Ito, H. Satoh, S. Okabe, Development and characterization of the partial nitrification aerobic granules in a sequencing batch airlift reactor, Bioresour. Technol. 139 (2013) 285–291. [14] A. Bhattacharya, A. Gupta, A. Kaur, D. Malik, Efficacy of Acinetobacter sp. B9 for simultaneous removal of phenol and hexavalent chromium from co-contaminated system, Appl. Microbiol. Biotechnol. (2014), http://dx.doi.org/10.1007/s00253-014-5910-5. [15] B. Liang, H.Y. Cheng, J.D.V. Nostrand, J.C. Ma, H. Yu, D.Y. Kong, W.Z. Liu, N.Q. Ren, L.Y. Wu, A.J. Wang, D.J. Lee, J.Z. Zhou, Microbial community structure and function of nitrobenzene reduction biocathode in response to carbon source switchover, Water Res. 54 (2014) 137–148. [16] I. Pala-Ozkok, A. Rehman, N. Yagci, E. Ubay-Cokgor, D. Jonas, D. Orhon, Characteristics of mixed microbial culture at different sludge ages: effect on variable kinetics for substrate utilization, Bioresour. Technol. 126 (2012) 274–282. [17] E. Dulekgurgen, N. Artan, D. Orhon, P.A. Wilderer, How does shear affect aggregation in granular sludge sequencing batch reactors? Relations between shear hydrophobicity and extracellular polymeric substances, Water Sci. Technol. 58 (2008) 267–276. [18] N.H. Rosman, A.N. Anuar, I. Othman, H. Harun, M.Z. Sulong, S.H. Elias, M.A.H.M. Hassan, S. Chelliapan, Z. Ujang, Cultivation of aerobic granular sludge for rubber wastewater treatment, Bioresour. Technol. 129 (2013) 620–623. [19] K.Z. Su, H.Q. Yu, Formation and characterization of aerobic granules in a sequencing batch reactor treating soybean-processing wastewater, Environ. Sci. Technol. 39 (2005) 2818–2827. [20] S.S. Adav, D.J. Lee, K.Y. Show, J.H. Tay, Aerobic granular sludge: recent advances, Biotechnol. Adv. 26 (2008) 411–423. [21] H.L. Wang, L. Li, P. Li, H. Li, G.S. Liu, J.M. Yao, The acceleration of sludge granulation using the chlamydospores of Phanerochaete sp. HSD, J. Hazard. Mater. 192 (2011) 963–969.
[22] Y. Liu, J.H. Tay, State of the art of biogranulation technology for wastewater treatment, Biotechnol. Adv. 22 (2004) 533–563. [23] B.S. McSwain, R.L. Irvine, P.A. Wilderer, The influence of settling time on the formation of aerobic granules, Water Sci. Technol. 50 (2004) 195–202. [24] K.V. Padoley, A.S. Rajvaidya, T.V. Subbarao, R.A. Pandey, Biodegradation of pyridine in a completely mixed activated sludge process, Bioresour. Technol. 97 (2006) 1225–1236. [25] G. Liu, Z. Ye, H. Li, R. Che, L. Cui, Biological treatment of hexanitrostilbene (HNS) produced wastewater using an anaerobic–aerobic immobilized microbial system, Chem. Eng. J. 213 (2012) 118–124. [26] D.H. Wen, J. Zhang, R.L. Xiong, R. Liu, L.J. Chen, Bioaugmentation with a pyridine-degrading bacterium in a membrane bioreactor treating pharmaceutical wastewater, J. Environ. Sci. 25 (2013) 2265–2271. [27] Y.H. Bai, Q.H. Sun, R.H. Sun, D.H. Wen, X.Y. Tang, Bioaugmentation and adsorption treatment of coking wastewater containing pyridine and quinoline using zeolite-biological aerated filters, Environ. Sci. Technol. 45 (2011) 1940–1948. [28] A.F. Duque, V.S. Bessa, M.F. Carvalho, M.K. de Kreuk, M.C.M. van Loosdrecht, P.M.L. Castro, 2-Fluorophenol degradation by aerobic granular sludge in a sequencing batch reactor, Water Res. 45 (2011) 6745–6752. [29] Y.H. Bai, Q.H. Sun, C. Zhao, D.H. Wen, X.Y. Tang, Microbial degradation and metabolic pathway of pyridine by a Paracoccus sp. strain BW001, Biodegradation 19 (2008) 915–926. [30] L. Qiao, D.H. Wen, J.L. Wang, Biodegradation of pyridine by Paracoccus sp. KT-5 immobilized on bamboo-based activated carbon, Bioresour. Technol. 101 (2010) 5229–5234. [31] L. Qiao, J.L. Wang, Microbial degradation of pyridine by Paracoccus sp. isolated from contaminated soil, J. Hazard. Mater. 176 (2010) 220–225. [32] Y. Lv, C. Wan, X. Liu, Y. Zhang, D.J. Lee, J.H. Tay, Freezing of aerobic granules for storage and subsequent recovery, J. Taiwan Inst. Chem. Eng. 44 (2013) 770–773.
