Aerobic granulation for 2,4-dichlorophenol biodegradation in a sequencing batch reactor

Chemosphere 69 (2007) 769–775 www.elsevier.com/locate/chemosphere

Aerobic granulation for 2,4-dichlorophenol biodegradation in a sequencing batch reactor Shu-Guang Wang

b

a,*

, Xian-Wei Liu a, Hua-Yong Zhang b, Wen-Xin Gong a, Xue-Fei Sun a, Bao-Yu Gao a

a School of Environmental Science and Engineering, Shandong University, Jinan 250100, China Energy and Environmental Research Center,North China Electric Power University, Beijing 102206, China

Received 15 November 2006; received in revised form 26 April 2007; accepted 3 May 2007 Available online 6 July 2007

Abstract Development of aerobic granules for the biological degradation of 2,4-dichlorophenol (2,4-DCP) in a sequencing batch reactor was reported. A key strategy was involving the addition of glucose as a co-substrate and step increase in influent 2,4-DCP concentration. After operation of 39 d, stable granules with a diameter range of 1–2 mm and a clearly defined shape and appearance were obtained. After granulation, the effluent 2,4-DCP and chemical oxygen demand concentrations were 4.8 mg l1 and 41 mg l1, with high removal efficiencies of 94% and 95%, respectively. Specific 2,4-DCP biodegradation rates in the granules followed the Haldane model for substrate inhibition, and peaked at 39.6 mg 2,4-DCP g1 VSS1 h1 at a 2,4-DCP concentration of 105 mg l1. Efficient degradation of 2,4-DCP by the aerobic granules suggests their potential application in the treatment of industrial wastewater containing chlorophenols and other inhibitory chemicals.  2007 Elsevier Ltd. All rights reserved. Keywords: 2,4-dichlorophenol (2,4-DCP); Aerobic granule; Biodegradation; Cometabolism; Haldane kinetics; Sequencing batch reactor (SBR)

1. Introduction Chlorophenols are xenobiotic contaminants that are often found in waste discharges of many industries including petrochemical, oil refinery, plastic, pesticides, biocides, wood preservers, pulp and insulation materials (Wang et al., 2000; Quan et al., 2004; Zilouei et al., 2006). Due to their high toxicity, strong odor emission and persistence in the environment and suspected carcinogenicity and mutagenity to living organisms, chlorophenols pose a serious ecological problem as environmental pollutants (Quan et al., 2004). Some chlorophenols have been listed as prior* Corresponding author. Tel.: +86 531 88362802; fax: +86 531 88364513. E-mail addresses: [email protected] (S.-G. Wang), xianweiliu00@ gmail.com (X.-W. Liu).

0045-6535/$ - see front matter  2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.05.026

ity pollutants (Zilouei et al., 2006). Hence, the removal of chlorinated phenols from wastewater is a necessary task to conserve the quality of natural water resources. Biological treatment can be a viable alternative for chlorophenol removal, since they could be mineralized by microorganisms under aerobic or anaerobic conditions (Zhang and Wiegel, 1990). A large amount of work has been carried out on chlorophenol degradation, particularly on aspects relating to different pure bacteria and fungi (Kim and Hao, 1999; Wang et al., 2000; Wang and Loh, 2001; Farrell and Quilty, 2002; Leontievsky et al., 2002; Kargi and Eker, 2005). However, information on aerobic chlorophenol degradation in bioreactor systems is still limited (Quan et al., 2004; Zilouei et al., 2006; Eker and Kargi, 2006), probably due to the toxicity or inhibition of chlorophenols to microorganisms (Quan et al., 2004). A few studies regarding bioreactor reported experimental

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results of employing immobilized biomass strategies such as attached biofilm to overcome the acute toxicity effect of chlorophenols. Compared with suspended systems, immobilized systems are able to retain a higher amount of chlorophenol-degrading bacteria, therefore achieving higher chlorophenol degradation activity and higher tolerance to chlorophenol toxicity (Quan et al., 2004; Eker and Kargi, 2006). Aerobic granulation represents an innovative cell immobilization strategy in biological wastewater treatment and it is attracting increasing interests (Beun et al., 1999; Beun et al., 2002; Zheng et al., 2005; de Kreuk and van Loosdrecht, 2006; Zheng et al., 2006). Aerobic granules are selfimmobilized microbial aggregates that are cultivated in sequencing batch reactors (SBRs) without adding a carrier material. Because of their compact structures, usage of bulky settling devices will be not needed (Beun et al., 1999; Jiang et al., 2002; Liu et al., 2006). Granulation facilitates the accumulation of high amounts of active biomass and the separation of this biomass from the wastewater liquor. The aggregation of microorganisms into compact aerobic granules also has additional benefits such as protection against predation and resistance to chemical toxicity (Jiang et al., 2004). Previous studies on aerobic granulation involved the use of readily biodegradable substrates such as ethanol, sucrose, and acetate (Beun et al., 1999; Beun et al., 2002; Zheng et al., 2005). Several recent studies reported the successful cultivation of aerobic granular sludge using toxic phenol and p-nitrophenol as substrates respectively (Jiang et al., 2004; Yi et al., 2006). In the case of chlorophenol, the presence of chloride in the ring makes it more challenging to degrade biologically. The chloro atom existing in the ring could inhibit the enzyme activity of ring cleavage, and thus enhance the resistance to biodegradation of chlorophenols (Wang and Werner, 2003). Chen et al. (2006) investigated the response of activated sludge to the presence of 2,4-dichlorophenol (2,4-DCP). It was demonstrated that at 1–20 mg l1, 2,4DCP slightly reduced the specific oxygen uptake rate of the activated sludge. The sludge yield slightly decreased with an increase in 2,4-DCP concentration, and its hydrophobicity also changed. Therefore, the main objective of this study was to explore the feasibility of cultivating aerobic granules for 2,4-DCP biodegradation, in order to develop aerobic granular sludge that can effectively treat 2,4-DCP and other xenobiotics. The process reported in this study can potentially be applied to remove 2,4-DCP in industrial effluents and contaminated groundwater. 2. Materials and methods 2.1. Reactor set-up and operation A 4-l column-type SBR (100 cm in height and 8 cm in diameter) used for the experiment was housed in a temperature-control room at 25 ± 2 C and operated sequentially in

a 4-h cycle, with 4 min of influent filling, 30 min of anoxic (no stirring), 200–210 min aeration, 1–11 min settling, and 5 min effluent withdrawal. Fine air bubbles for aeration were supplied through a dispenser at the reactor bottom at a superficial gas velocity of 2.76 cm s1. The influent entered through the reactor bottom. Effluent was discharged at a volumetric exchange ratio of 50%, giving a hydraulic retention time (HRT) of 8 h. The abiotic loss of 2,4-DCP in SBR was negligible under identical operating conditions. 2.2. Seed sludge and wastewater The activated sludge was obtained from the secondary clarifier of the Jinan No. 2 Municipal Wastewater Treatment Plant, China, and it was conditioned in an aeration tank over a 30-d acclimation period to allow the biomass adapt to 2,4-DCP. During this acclimation period, the aeration tank was fed in batch with a synthetic wastewater containing 1000 mg l1 of glucose and 2,4-DCP, stepwise increased from 10 to 50 mg l1. After acclimation, 2 l of acclimated sludge were inoculated into the SBR, resulting in a mixed liquor suspended solids (MLSS) concentration of 4.9 g l1 in the SBR. The seeding sludge was grayish brown and its structure was fluffy, irregular and loose (Fig. 1a). The SBR was fed with a synthetic wastewater containing 1000 mg l1 glucose and 2,4-DCP, and its concentration was gradually increased from 50 mg l1 to 100 mg l1. The synthetic wastewater consisted of a buffered mineral salt medium with the following composition (mg l1): Na2HPO4, 50; KH2PO4, 50; Ca (NO3)2, 30; MgSO4 Æ 7H2O, 25; FeSO4 Æ 7H2O, 20. Micronutrients were supplemented to the medium at 1 ml l1 (Moy et al., 2002). Urea with the concentration of 50 mg l1 was used as nitrogen source. 2.3. Batch biodegradation kinetics The ability of granules to degrade 2,4-DCP was evaluated in 250-ml serum vials which contained the nutrient medium mentioned above and the target contaminant at predetermined concentrations ranging from 50 to 300 mg l1. For each batch, 100 ml of aerobic granules (equivalent to 0.49 g VSS) was harvested from the SBR at the end of aeration period, washed twice, and added into the bottle. Vials were then shaken at 25 C on an orbital shaker at 200 rpm for 12 h and assayed periodically. All experiments were performed in triplicate. A kinetic analysis of the degradation data was performed on the basis of Haldane’s equation for describing biodegradation of an inhibitory substrate (Yi et al., 2006) V ¼

V max S 2

K s þ S þ SK i

;

ð1Þ

where V and Vmax are the specific and the maximum specific substrate degradation rates (mg 2,4-DCP g1 VSS1 h1), respectively, and S, Ks and Ki are the substrate

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3. Results and discussion 3.1. Formation of aerobic 2,4-DCP-degrading granules

2.4. Analytical methods Measurement of pH, biomass concentration, chemical oxygen demand (COD) were conducted in accordance with the Standard Methods (APHA, 1998). The 2,4-DCP concentration was determined spectrophotometrically using the absorbance values at 306 nm with a UV/vis spectrophotometer according to Wu and Yu (2007). Inorganic chloride was determined with an ion-chromatograph (DX-100 Ion chromatography, Dionex, USA) equipped with an Ion Pac As14 column (4 mm · 250 mm). Morphology and surface structure of granules were observed qualitatively with a scanning electron microscope (SEM) (Hitachi S-570, Japan). Granules were prepared for SEM image by washing with a phosphate buffer and fixing with 2% glutaraldehyde overnight at 4 C. Fixed granules were washed with 0.10 M sodium cacodylate buffer, dehydrate by successive passages through 25, 50, 75, 80, 90, 95 and 100% ethanol and dried with a CO2 Critical Point Dryer.

3.2. Performance of the reactor The influent 2,4-DCP concentrations persisted at levels of approximately 50 mg l1 during the initial 23 d (Fig. 3a). The effluent 2,4-DCP concentration was 6.2 mg l1 on day 4 of the granule cultivation phase, and 6

-1

concentration, half-saturation constant, and inhibition constant (mg l1), respectively.

MLSS (g l )

Fig. 1. Sludge morphology: (a) seed sludge and (b) granular sludge (42 days after inoculation).

Activated sludge from a municipal wastewater treatment plant was initially acclimated over a 30-d period to allow the biomass to adapt to 2,4-DCP. The initial sludge had a MLSS value of 2.5 g l1, while the MLSS concentration was increased to 4.9 g l1 after the acclimation period (Fig. 2). The acclimated sludge was used as seed sludge for the cultivation of aerobic 2,4-DCP-degrading granules. Aerobic granules were first observed on day 8 as small spherical particles were dispersed with the amorphous sludge flocs. However, the biomass concentration decreased from 4.9 to 2.3 g l1 attributed to the reduced settling time from 11 to 5 min on day 5. The biomass with the settling velocities smaller than 6 m h1 was washed out of the reactor gradually. The small granules grew rapidly in the subsequent weeks, while more floc-like sludge was washed out, resulting in the accumulation of granules. On day 22, settling time was decreased from 5 min to 1 min. Therefore, the selective pressure led to the wash out of biomass, meanwhile enhanced the granules to grow to become the dominant form of biomass in the reactor, as evidenced by the gradual increase in biomass concentration beyond day 23 (Fig. 2). The rapid transition from acclimated activated sludge to aerobic granules shows that the operating strategy chosen was effective. The biomass concentration in the reactor generally showed an upward trend to stabilize at 4.7 g of MLSS l1 from day 39 (Fig. 2). Stable granules were obtained in the SBR (Fig. 1b). The diameter of the aerobic granules was in the range of 1–2 mm.

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Time (day) Fig. 2. Profiles of MLSS.  MLSS. Arrows indicate the changeds in settling time.

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2,4-DCPCon.(mg l )

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Time (day) Fig. 3. Performance for the continuous operation of SBR with respect to: (a) 2,4-DCP and (b) COD. d 2,4-DCP of Inf.; s 2,4-DCP of Eff.;  COD of Inf.;  COD of Eff. and m Removal of Efficiency.

the acclimated sludge showed a marked improvement in 2,4-DCP degradation ability. However, the effluent 2,4DCP concentrations were more than 20 mg l1 due to the wash out of the poor-settling flocs on day 5 when the selective pressure was applied. With the increase in biomass concentration and the microbial adaptation to 2,4-DCP, the effluent 2,4-DCP concentration gradually decreased to 5.9 mg l1 on day 23, resulting in a removal efficiency of 88%. Afterwards, the influent 2,4-DCP concentration was increased stepwise from 50 to 100 mg l1 from day 25 to the end of the operation. The increase in influent 2,4DCP concentration did not diminish the ability of the granules to remove 2,4-DCP. The average effluent 2,4-DCP concentration and removal efficiency were stabilized at less than 4.8 mg l1 and above 94%, respectively. The influent COD concentrations slightly fluctuated around 920 mg l1, which resulted in an organic loading rate of 2.8 kg COD m3 d1 (Fig. 3b). The average effluent COD concentration and COD removal efficiency reached 41 mg l1 and 95%, respectively, after granulation. Glucose in the synthetic wastewater mainly contributed to the COD in the influent. However, the removal characteristic of COD was in good accordance with that of 2,4-DCP and the increase in influent 2,4-DCP concentration had no effect on the COD removal efficiency in the reactor. Since the influent contained both glucose and 2,4-DCP, the similar removal characteristics suggest that glucose was not preferentially degraded at the expense of 2,4-DCP, when the benign substrates was supplied together with 2,4-DCP.

A key strategy in the cultivation of stable aerobic 2,4DCP-degrading granules was dose of glucose as a co-substrate to promote granule formation. This was different from previous studies, in which target contaminant was used as a sole carbon and energy source to develop aerobic granules for biodegradation (Jiang et al., 2002; Tay et al., 2005). Similar strategies have been adopted in chlorophenol-degrading biofilm or granule-based reactors to facilitate the biofilm adhesion and co-metabolic transformation of the contaminant. For example, Eker and Kargi (2006) added molasses along with 2,4-DCP to ensure an acclimation of the attached biofilm to 2,4-DCP. Hendriksen et al. (1992) reported that the addition of glucose was beneficial to the substrate utilization, stimulated the dechlorination, and maintained sufficient biomass in an upflow anaerobic sludge bed reactor treating pentacholophenol. Bali and S ß engu¨l (2002) reported that in co-metabolic transformation of 4-chlorophenol, phenol is a good primary substrate since it can not only easily induce the monooxygenase required for 4-chlorophenol transformation, phenol oxidation can also efficiently regenerate consumed NADH. Strong competitive inhibition between phenol and 4-chlorophenol, however, inhibits 4-chlorophenol transformation significantly. It has been found that 4-chlorophenol was transformed rapidly only after phenol was almost fully depleted (Loh and Wang, 1998). In this case, the co-metabolic enzymes required for 4-chlorophenol transformation were most probably induced by 4-chlorophenol. As for the NADH required for 4-chlorophenol transformation, this could be efficiently formed through the oxidation of glucose (Bali and S ß engu¨l, 2002). With the rapid oxidation of glucose, NADH is quickly regenerated, consequently facilitating the transformation of 4-chlorophenol. Comparison our study with the previous literature above-mentioned, it can be inferred that in our study, the addition of glucose, increased the utilization rate of co-substrates and accelerated NADH regeneration, which consequently facilitated aerobic granulation and then the transformation of 2,4-DCP. However, this warrants a further investigation. 3.3. 2,4-DCP degradation in one cycle The removal of 2,4-DCP and the release of chloride in one cycle of the SBR operation were examined on days 27 and 43, respectively (Fig. 4). On day 27, the 2,4-DCP concentration was decreased from 35 mg l1 to 7.6 mg l1 in 150 min in one cycle and became stabilized at 7.6 mg l1 afterwards. Correspondingly, the chloro concentration released increased to 10.2 mg l1 after 150-min operation. On day 43, the 2,4-DCP concentration increased to 48 mg l1 at the beginning of the SBR cycle. However, 2,4-DCP reached 10.4 mg l1 in 90 min and the chloro concentration released increased to 14.4 mg l1 at the same time. The degrading rate of 2,4-DCP on day 43 was higher than that on day 27, attributed to the greater biomass concentration and higher biomass activity on day 43.

2, 4-DCP Conc. (mg l-1)

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Time course of 2,4-DCP concentrations is shown in Fig. 5. 2,4-DCP concentration in the batch reactors decreased with time steeply at a lower initial 2,4-DCP content, as compared to those containing higher levels of 2,4DCP. The 2,4-DCP contents of the control reactors did not change significantly (<3%) during the course of experiments. The extent of adsorption and non-biological degradation of 2,4-DCP was negligible. Fig. 6 illustrates the variation of initial 2,4-DCP degradation rate with its initial concentration. The specific 2,4DCP degradation rate increased with the increasing initial 2,4-DCP content up to 105 mg l1, peaked at 39.6 mg 2,4DCP g1 VSS1 h1, but then decreased for the initial 2,4-DCP contents between 105 and 300 mg l1, indicating the inhibitory effect of 2,4-DCP at a concentration exceeding 105 mg l1. Therefore, the Haldane equation was used to model the degradation data. The kinetic parameters estimated with the least-square error method were Vmax = 478 mg 2,4-DCP g1 VSS1 h1, Ki = 19.1 mg l1, and Ks = 587 mg l1, with a correlation coefficient (R2) of 0.999. Therefore, the resulting kinetic equation is as below:

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Time(min) Fig. 4. Typical concentration profiles during a cycle of the SBR measured at day 27 (a) and day 43 (b).  2,4-DCP; m Cl1.

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V ¼

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ð2Þ

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The ability of the aerobic granules to degrade 2,4-DCP was evaluated by monitoring 2,4-DCP disappearance at its different concentrations in batch culture on day 45.

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Time (h) Fig. 5. Time courses of 2,4-DCP concentration for different 2,4-DCP contents.  50 mg l1;  100 mg l1; m 150 mg l1; n 200 mg l1; s 250 mg l1; d 300 mg l1.

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3.4. 2,4-DCP biodegradation kinetics

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Quan et al. (2004) reported that the biosorption of 2,4DCP onto activated sludge occurred in the initial 10 h in a bioaugumented activated sludge system. However, in the present study, an increase in chloro concentration was found, and was stoichiometrically related to the amount of 2,4-DCP degraded. This suggests that 2,4DCP was soundly degraded by the aerobic granules. The critical step in the degradation of chlorinated compounds is the cleavage of the carbon–chlorine bond (Mohn and Kennedy, 1992; Fahmy et al., 1994; Snyder et al., 2006). Two different catabolic pathways appear to have evolved for chlorophenol degradation, which involves dehalogenation after ring cleavage and dehalogenation before ring cleavage (Snyder et al., 2006). In the case of 2,4-DCP degradation, the first step is the production of 3,5-dichlorocatechol with the use of 2,4-dichlorophenol hydroxylase. Later, 3,5-dichlorocatechol is converted to 2,4-dichlorocis,cis-muconate, which is catalyzed by 3,5-dichlorocatechol-1,2-dioxygenase (Radjendirane et al., 1991), and is then ultimately mineralized to CO2 in some cases. Therefore, specific enzyme induced and its distribution with cells would play a key role in the reductive dechlorination after ring cleavage.

2,4-DCP Con.(mg l-1)

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2,4-DCPCon. (mg l-1) Fig. 6. Specific 2,4-DCP degradation rates of the aerobic granules at different 2,4-DCP concentrations.

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Sahinkaya and Dilek (2007) reported the parameters of Vmax, Ki and Ks were 112 mg 2,4-DCP g1 VSS1 h1, 44.46 mg l1 and 13 mg l1, respectively, in a floc-based SBR. In the study involving aerobic granule-based SBR for phenol removal by Jiang et al. (2002), the kinetic parameters were found to be 234 mg phenol g1 VSS1 h1, 212 mg l1, 481 mg l1, respectively. The flocs reported by Sahinkaya and Dilek (2007) have a very high rate at low concentrations, but at higher concentrations they are very rapidly inhibited, resulting in similar rates as the granules in this study at high 2,4-DCP concentrations. The granules from Jiang et al. (2002) seem to be protected for the phenol or inhibition starts at very high concentrations, since the rate keeps the increasing trend until 240 mg l1. In the granules specific organisms can accumulate, since they are not washed out of the system as long as they are incorporated in the granule. The solid retention time (SRT) inside the granules is much higher than one can reach in a flocculated sludge system, so maybe the specific organisms that can degrade or are adapted to 2,4-DCP need this long SRT. 4. Conclusions This study demonstrates that compact well-setting aerobic granules could be successfully cultivated in an SBR for 2,4-DCP biodegradation using a strategy involving the addition of glucose as a co-substrate and step increase in the influent 2,4-DCP concentration. After 39-d operation, stable granules with a diameter of 1–2 mm were obtained. After granulation, the average effluent 2,4-DCP and COD concentrations were 4.8 mg l1 and 41 mg l1, respectively, and corresponding removal efficiencies of 94% and 95%, respectively. Specific 2,4-DCP biodegradation rates in the granules followed the Haldane model, and peaked at 39.6 mg 2,4-DCP g1 VSS1 h1 at a 2,4-DCP concentration of 105 mg l1. Therefore, this study demonstrate the technical feasibility of utilization of aerobic granules for the effective degradation xenobiotic contaminants like 2,4DCP. Acknowledgement Authors wish to thank the Natural Science Foundation of Shandong Province, China (Grant No. Q2005B01) for the support on this work. References APHA, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, Washington, DC. Bali, U., S ß engu¨l, F., 2002. Performance of a fed-batch reactor treating a wastewater containing 4-chlorophenol. Process Biochem. 37, 1317– 1323. Beun, J.J., Hendriks, A., van Loosdrecht, M.C.M., Morgenroth, E., Wilderer, P.A., Heijnen, J.J., 1999. Aerobic granulation in a sequencing batch reactor. Water Res. 33, 2283–2290.

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