Use of submerged anaerobic–anoxic–oxic membrane bioreactor to treat highly toxic coke wastewater with complete sludge retention

Journal of Membrane Science 330 (2009) 57–64

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Use of submerged anaerobic–anoxic–oxic membrane bioreactor to treat highly toxic coke wastewater with complete sludge retention Wen-Tao Zhao a , Xia Huang a,∗ , Duu-Jong Lee b , Xiao-Hui Wang a , Yue-Xiao Shen a a b

State Key Joint Laboratory of Environment Simulation and Pollution Control, Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, China Chemical Engineering Department, National Taiwan University, Taipei 106, Taiwan

a r t i c l e

i n f o

Article history: Received 5 November 2008 Received in revised form 17 December 2008 Accepted 18 December 2008 Available online 19 January 2009 Keywords: Membrane bioreactor Complete sludge retention Coke wastewater Microbial characteristics Membrane filterability

a b s t r a c t Coke wastewater is an extremely toxic industrial effluent that requires treatment before discharge. A bench-scale, anaerobic–anoxic–oxic membrane bioreactor (A1 /A2 /O-MBR) system was utilized to treat real coke wastewater with complete sludge retention. In a 160-d test, the A1 /A2 /O-MBR system stably removed 87.9 ± 1.6% of chemical oxygen demand, 99.4 ± 0.3% of turbidity, and 99.7 ± 3.5% of NH4 + -N from coke wastewater. The membrane rejected almost all suspended solids; hence, a low food-to-microorganism environment was created to degrade refractory substances and reduce sludge production rates. The microbial diversity in the MBR system declined over time; however, neither pollutant removal efficiency nor total biological activity was adversely affected. Membrane fouling, which occurred during the operation of the MBR system, was principally resulted from the colloidal fraction of supernatant in suspension. Physical cleaning removed initial deposits of particles; however, prolonged operation resulted in severe clogging that can only be removed by chemical cleaning. An A1 /A2 /O-MBR system with short intermittent physical cleaning was recommended for coke wastewater treatment. © 2009 Elsevier B.V. All rights reserved.

1. Introduction Membrane bioreactors (MBRs) comprise a promising wastewater treatment technology that is widely applied in treating municipal and industrial wastewaters [1–6]. By effective biomasseffluent separation with microfiltration or ultrafiltration membrane modules, an MBR can achieve complete sludge retention for attaining high-sludge concentrations and long sludge retention times (SRTs). In other words, MBRs are operated at very low food-to-microorganism (F/M) ratios for complete mineralization of biodegradable organic matter and low-net sludge production rates [7–9]. Slow-growing microorganisms, such as nitrifying bacteria, can be enriched in this complete retention system for achieving high and stable nitrification efficiency [10,11]. Moreover, efficient biological nitrogen removal can be achieved when an MBR is integrated with anoxic zones [11,12]. Coke wastewater, one of the most toxic industrial effluents, is generated from coal carbonization and fuel classification processes in the iron and steel industry [13]. Typical coke wastewater contains high-strength ammonia, cyanide, thiocyanate and phenols, as well as certain quantities of poly-nuclear aromatic hydrocarbons (PAHs)

∗ Corresponding author. Tel.: +86 10 6277 2324; fax: +86 10 6277 1472. E-mail address: [email protected] (X. Huang). 0376-7388/$ – see front matter © 2009 Elsevier B.V. All rights reserved. doi:10.1016/j.memsci.2008.12.072

and nitrogen, oxygen or sulfur-containing heterocyclic compounds [14–16]. Coke wastewater is conventionally first subjected to a physico-chemical pre-treatment, such as phenol extraction, ammonia stripping, or gas flotation, to reduce contaminant loads, followed by a polishing stage using biological treatments [17]. Among various proposed biological processes, the anaerobic–anoxic–oxic (A1 /A2 /O) system and anoxic–oxic (A/O) system preferentially remove nitrogen, cyanide, thiocyanate, phenols, and polycyclic and heterocyclic compounds [16,18–21]. However, stability of these biological treatment stages is generally poor as coke wastewater is available intermittently [13,16]. Additionally, surplus sludge from a biological treatment stage is typically contaminated with high levels of residual PAHs [22]. An integrated (A1 /A2 /O-MBR) system with complete sludge retention is a promising scheme for treating highly toxic coke wastewater. Many studies have analyzed long-term MBR operation with synthetic or municipal wastewater without sludge discharge [8,9,11,23–25]. Few MBR studies focused on long-term treatability of industrial effluents with high-toxic loadings. This study explores long-term organic- and nitrogen-removal performance of a bench-scale anaerobic–anoxic–oxic submerged MBR (A1 /A2 /OMBR) system for real coke wastewater treatment with complete sludge retention. The microbial community characteristics, suspension characteristics and membrane fouling potentials were determined.

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Fig. 1. Schematic diagram of A1 /A2 /O-MBR system.

2. Materials and methods 2.1. The reactors The bench-scale A1 /A2 /O-MBR system (Fig. 1) consisted of an anaerobic reactor (A1 , working volume of 11.5 L, packed with a soft medium of hydrophilic acrylate fibre at 95% porosity), an anoxic reactor (A2 , working volume of 12.0 L, completely mixed), an oxic reactor (O, working volume of 27.0 L, completely mixed), and a submerged hollow fibre polythene membrane (nominal pore size, 0.4 ␮m; membrane area, 0.2 m2 ) (Mitsubishi Rayon Co. Ltd., Japan). A diaphragm pump (model X068-XB-AAAA365, Pulsafeeder, USA) fed the raw wastewater from a 100-L storage tank into anaerobic reactor A1 . The internal mixed liquor recirculation flow from O to A2 was 300% of the oxic outlet flow, and the recirculation was performed by a peristaltic pump (model no. 7520-57, Cole-Palmer, USA). Air diffusers underneath the membrane in the oxic reactor provided a constant air flow of 2 m3 /h to mix the suspensions, scored the membrane surface for fouling reduction and maintained a dissolved oxygen (DO) concentration of >3 mg/L throughout the 160-d test. The permeate was drawn at a filtration/idle cleaning ratio of 8 min/2 min by a suction pump in constant-rate mode (model X068-XB-AAAA365, Pulsafeeder, USA). The Na2 CO3 solution was added to compensate for alkalinity loss in the nitrification reaction and maintain pH in the oxic reactor at 7.0–7.2 using a pH controller (model Liquitron DP 5000, LMI MILTON ROY, USA) and an electromagnetic dosing pump (model UL#A752-393SI, LMI MILTON ROY, USA). The temperatures of anaerobic (A1 ), anoxic (A2 ) and oxic reactors (O) were maintained at 30–35 ◦ C using thermostats (this temperature range was close to that in coke wastewater treatment facilities). A control A1 /A2 /O-CAS system, comprising an A1 reactor (working volume of 11.5 L, packed with a soft media of hydrophilic acrylate fibre at 95% porosity), an A2 reactor (working volume of 12.0 L, completely mixed), an O reactor (working volume of 27.0 L, completely mixed), and a settling tank (working volume of 13.5 L) had the same working volumes, temperature, DO content, and internal circulation ratios as the test A1 /A2 /O-MBR system. 2.2. Wastewater and tests Raw coke wastewater was collected bi-weekly from the Beijing Steel Company, China, and stored at 4 ◦ C prior to tests. The wastewater was pretreated using ammonia stripping and gas flotation in the coke factory to decrease ammonia and oil concentrations. Phos-

phoric acid was added to adjust the pH to neutral and act as a phosphorus source for microorganisms. The A1 reactor was first seeded with sludge collected from a bench-scale sequential batch reactor (SBR), and the mixed liquor suspended solids (MLSS) concentration after inoculation was 15.0 g/L. The anaerobic reactor was operated using coke wastewater for 2 additional months at a hydraulic retention time (HRT) of 8.5 h. The A1 effluent was then fed into the A2 /O reactors, which were seeded with 5.8 g-MLSS/L sludge collected from the return sludge stream in a full-scale A2 /O treatment plant for coke wastewater. Finally, the A1 /A2 /O system was inserted with a submerged membrane module for a 160-d test. The HRT of A1 reactor was first set at 8.5 h, and reduced to 4.3 h by adjusting the working volume after 90-d of testing. The HRTs for the A2 and O reactors were always 8.9 and 20.0 h. The average membrane flux was 6.75 L/(m2 h) and total flow rate was 1.35 L/h. Except for sampling and membrane cleaning, no sludge was discharged from the reactors; the quantity of sludge discharged was minimal. Ex situ physical cleaning (jetting the module with tap water under moderate pressure) and chemical cleaning (soaking the module in 0.05% NaClO solution for 24 h) were applied to restore membrane permeability when needed. Membrane permeability was determined using 25 ◦ C tap water right after cleaning; the sodefined permeability is called standard permeability. The standard permeability of a new membrane is 11.0 L/(m2 h kPa). The control A1 /A2 /O-CAS system was started up simultaneously with A1 /A2 /O-MBR system with the same operational conditions except for SRT. The average SRT for reactors A2 and O was 100 d. 2.3. Sampling and analysis Influent wastewater, anaerobic effluent, CAS effluent and membrane permeate samples were collected from the feed tank, A1 reactor outlet, settling tank outlet and effluent pump outlet from the MBR, respectively. Anoxic and oxic supernatant samples were collected and extracted by centrifugation at 4000 rpm for 10 min, followed by filtration through a 0.45 ␮m membrane. Anoxic and oxic sludge samples for concentration measurement were collected from each reactor. The tests of microbial characteristics were conducted on samples collected from the oxic reactor. Biochemical oxygen demand (BOD5 ), chemical oxygen demand (COD), NH4 + -N, total nitrogen (TN), nitrite, MLSS and mixed liquor volatile suspended solids (MLVSS) for collected samples were measured using standard methods [26]. Sample turbidity was monitored by a digital direct-reading turbidimeter (model 965-10, Orbeco-Hellige, USA). The DO and pH of suspensions were recorded

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Fig. 2. COD removal in the A1 /A2 /O-MBR system.

by a dissolved oxygen meter (model 850, Thermo Orion, USA) and pH meters (model 868, Thermo Orion, USA), respectively. Specific oxygen uptake rate (sOUR) was measured relative to substrate oxidation rate [27] using phenol, ammonium chloride and sodium nitrite as carbonaceous and nitrogenous substrates, respectively. The sludge floc was examined using a conventional optical microscope (XSZ-H, Chongqing Optical & Electrical Instrument Co. Ltd., China). Gram staining was applied to identify bacterial morphology. The bacterial community was analyzed using the terminal-restriction fragment length polymorphism (T-RFLP) method. In other words, the genomic DNA of sludge samples was extracted using a FastDNA® SPIN Kit for Soil (Bio101, Carlsbad, CA, USA) according to the manufacturer’s protocol. The 16S rRNA and amoA from the obtained DNA were subjected to PCR amplification in a thermocycler (GeneAmp PCR System 9700, Applied Biosystems, USA). The 16S rRNA gene was amplified using eubacterial universal primers 8F (5 -AGAGTTTGATCCTGGCTCAG-3 ) and 1492R (5 -GGT2TACCTTG TTACGACTT-3 ). For amoA gene amplification, primers amoA-1F (5 -GGGGTTTCTACTGGTGGT-3 ) and amoA-2R (5 -CCCCTCKGSA AAGCCTTCTTC-3 ; K is G or T; S is C or G) were utilized as described by Park et al. [28]. All PCR amplification was conducted with the forward primer 5 -labeled with the FAM dye. The amplified PCR products were then purified using a QIAquicks PCR purification kit (Qiagen, Valencia, CA, USA) and digested with Rsa I and Taq I restriction endonuclease (MBI Fermentas, Hanover, MD, USA) for the 16S rRNA gene and amoA gene, respectively, following the manufacturer’s instructions. The fluorescently labeled terminal restriction fragments (T-RFs) were run through an automated sequence analyzer (ABI Prism 3130-Avant genetic analyzer, Applied Biosystems, USA) in the GeneScan mode; the length of TRFs were determined by comparison with internal standards. The Shannon–Weaver diversity index (H), sample richness (S) and evenness (E) were calculated as described in [12]. The particle size distribution in sludge was determined by light scattering using a Malvern 2601LC Master instrument. Viscosity (a ) was analyzed using a rotational viscosity meter (model LVDV I, Brookfield, UK). Shear rate (dv/dy) was set at 16.7 s−1 for all sludge a measurements.

Total organic carbon (TOC) of oxic supernatant was roughly used as the indicator of soluble metabolic products concentration in reactor, which included the colloidal and dissolved matter. The corresponding colloidal organic carbon (COC) and dissolved organic carbon (DOC) fractions adhere to TOC = COC + DOC. The COC and DOC concentrations were determined using the method described by Bouhabila et al. [29]. Briefly, the supernatant, which contained colloidal and dissolved matter, was obtained by direct centrifugation (4500 rpm for 1 min). Flocculation with Al2 (SO4 )3 at 250 mg/L and a second centrifugation (4500 rpm for 10 min) of the supernatant removed all colloids. In this manner, two supernatant samples were available for TOC determination (TOC-5000A, SHIMADZU, Japan). The fouling potential of supernatant fractions was assessed by flat sheet membrane filtration tests using a stirred dead-end filtration cell (model 8400, Amicon Corp., USA) and hydrophilic microfiltration membranes with a pore size of 0.22 ␮m (GVWP, Millipore, USA). Filtration pressure and temperature were maintained at 5 kPa and 25 ◦ C, respectively. Membrane flux was determined by weighing the permeate on a top-loading balance. New membranes were used for each test. 3. Results and discussion 3.1. Degradation performance The A1 /A2 /O-MBR system was operated continuously for 160 d without withdrawing any sludge. Satisfactory COD biodegradation was achieved right after the A1 /A2 /O-MBR system was started. This satisfactory performance was maintained over the entire experimental period, regardless of the fluctuating influent COD concentration of 1400–3020 mg/L (Fig. 2). The oxic supernatant COD fluctuated at 250–630 mg/L; however, the permeate concentration from the MBR remained relatively stable. Increased loading occurred from the 121st to 133rd days and did not adversely affect effluent quality. Average effluent COD concentration was 243 ± 39 mg/L (n = 72); removal efficiency was 87.9 ± 1.6% (n = 72). At the same influent concentrations, the effluent COD in the pro-

Table 1 Comparison of the effluent pollutant concentrations between A1 /A2 /O-MBR and A1 /A2 /O-CAS systems. Values

Mean Range nb

COD (mg/L)

NO2 − -N (mg/L)

NH4 + -N (mg/L)

Turbidity (NTU)

TN (mg/L)

MBR

CAS

MBR

CAS

MBR

CAS

MBR

CAS

MBR

CAS

243 ± 39a 158–339 72

313 ± 50 230–517 72

1.8 ± 0.2 0.4–1.4 37

89 ± 4.0 23–187 37

0.8 ± 0.9 0.1–12.5 44

17.7 ± 9.9 0.5–84.7 44

4.2 ± 3.5 0.1–14.2 44

2.8 ± 2.8 0.1–24.5 44

168 ± 39 73–250 72

189 ± 35 104–316 72

NH4 + -N and NO2 − -N data used for statistical analysis were obtained from the 60th to 160th days. a Mean ± standard deviation. b Number of measurement.

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Table 2 The contribution of each stage to pollutant removal in the A1 /A2 /O-MBR system. Stage

COD (n = 29)a

A1 (%) A2 (%) O (%) M (%)

± ± ± ±

8.7 50.8 33.4 7.1

NH4 + -N (n = 31)

b

6.2 12.7 12.6 2.4

−1.3 −0.8 101.9 0.2

± ± ± ±

6.6 16.6 16.5 0.5

TN (n = 29) −4.1 108.3 −10.8 6.4

± ± ± ±

9.8 24.2 24.9 4.3

Sample measurements were performed about every 5 d during the whole operation period and the data were used to evaluate the contribution of each stage. A1 : anaerobic reactor; A2 : anoxic reactor; O: oxic reactor; M: membrane interception. a Number of measurement. b Mean ± standard deviation.

posed system was more stable and the mean concentration was 70 mg/L lower than that in the A1 /A2 /O-CAS system (Table 1). In the A1 /A2 /O-MBR system, the A1 reactor removed <10% of COD, whose role was primarily to convert refractory and inhibitory compounds into biodegradable organic substances [15,16]. Nitrification occurred principally in the oxic reactor, whereas anoxic reactor (A2 ) denitrified nitrate and removed >50% of COD from the entire A1 /A2 /O-MBR system (Table 2). The oxic reactor with membrane removed 40.5% of COD; membrane interception accounted for 7.1% of COD removal. Restated, the membrane retained sludge in the system, leading to a turbidity removal rate of 99.4 ± 0.3% (n = 37), while microorganisms in the oxic reactor degraded organic matter. The effluent BOD5 /COD ratio was <0.06, suggesting that most residual COD was highly refractory. The A1 /A2 /O-MBR system satisfactorily removed most biodegradable COD in the coke wastewater samples. The NH4 + -N was easily oxidized by the A1 /A2 /O-MBR system; nitrite accumulation was only notable within 4 d after system startup (Fig. 3). The removal rate of NH4 + -N was 99.7 ± 3.5% (n = 44) after 60 d of operation, likely due to the time needed to enrich slowgrowing autotrophic nitrifying bacteria for complete nitrification of ammonia [30]. Considering the effluent NH4 + -N concentrations (Table 1), the nitrification was more stable and complete in the A1 /A2 /O-MBR system than that in the control one. Moreover, the maximum TN removal rate was 74.9%, near the theoretical value of 75% with 300% of the recycling flow. However, mean daily removal rate fluctuated significantly at 50.2 ± 10.9% (n = 72) (Fig. 4); this may be attributed to the response to large variations in coke wastewater composition over time [31]. Generally, the anoxic unit accounted for most TN removal (Table 2). 3.2. Microbial growth and viability The analysis of microbial characteristics primarily focused on anoxic and oxic reactors, comprising a single sludge recycling stream between the two reactors, in terms of the following factors: (1) anoxic and oxic reactors (combined with a membrane to form an

MBR) were mainly used for pollutant removal, whereas the anaerobic biofilm reactor was used primarily as a pre-treatment unit for increasing the biodegradability of coke wastewater; and (2) complete biomass retention directly affected microbial behavior in the anoxic and oxic reactors and not the anaerobic reactor. Average sludge concentration in the anoxic and oxic reactors increased consistently during the first 40 d; the MLSS was then maintained at roughly 8.7 g/L for 50 d (Fig. 5). During the following 70 d, as influent COD and nitrogen loads increased (Figs. 2–4), sludge concentration increased and leveled off at approximately 15.0 g/L. The MLVSS/MLSS ratio during the entire 160-d operation was roughly 90.2 ± 1.0%, indicating that no inorganic matter accumulation occurred in the anoxic and oxic reactors. The MLVSS/MLSS ratio achieved in this study was slightly higher than those commonly reported for municipal wastewater (70–80%) [32]. This high ratio was likely attributable to the retention of inorganic matter by biofilms in reactor A1 . The corresponding food-to-microorganism (F/M) ratio in the anoxic and oxic reactor decreased and approached a mean value of 0.15 kg COD/(kg MLVSS d), generating a low-sludge yield of 0.035 kg MLVSS/kg COD (data not shown), which was onethird of that of the control system. The complete sludge retention could allow system operation at higher sludge concentrations, which resulted in lower F/M ratio. As a result, the pollutants were mainly utilized by microorganisms for non-growth energydemanding activities rather than biosynthesis, which favored the reduction of net sludge production [11,33]. The sludge yield by the A1 /A2 /O-MBR system was markedly lower than those reported for MBRs treating municipal wastewater (0.12 kg MLVSS/kg COD) [33], synthetic wastewater (0.11 kg MLVSS/kg COD) [8], and seafood wastewater (0.09 kg MLVSS/kg COD) [25]. The proposed system produced low quantities of sludge, likely due to the fact that some sludge could be transported by recirculation stream back to A2 reactor and was degraded there. Additionally, the sOUR fluctuated at 50–70 mg O2 /(gVSS h), indicating a roughly constant microbial activity in the 160-d test (data not shown). Considering the highsludge concentrations in the proposed system, high-total microbial viability could be achieved, thus ensuring the high-removal efficiencies of COD and NH4 + -N. 3.3. Microbial diversity Microscopic observations showed that immediately after MBR startup, protozoa and metazoa were in detectable quantities in the oxic reactor, but declined after 70 d of operation (data not shown). Consequently, neither protozoa nor metazoa was effective microorganisms for the MBR system. Filamentous bacteria never appeared throughout the study period; conversely, short rod-shaped bacteria were the predominant group in the sludge. This observation disagreed with those that detected filamentous bacteria when treating municipal wastewater by an MBR with complete biomass retention

Fig. 3. NH4 + -N removal in the A1 /A2 /O-MBR system.

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Fig. 4. TN removal in the A1 /A2 /O-MBR system.

[34]. The high toxicity of feed wastewater may hinder the development of filamentous bacteria. The absence of filamentous bacteria benefits MBR operation as excess growth of filamentous bacteria significantly increases suspension viscosity during membrane filtration [35,36]. The 16S rRNA and amoA genes were utilized to target total bacteria and autotrophic ammonia oxidizing bacteria, respectively. The bacterial T-RFLP fingerprints had similar fragment lengths, which mainly corresponded at T-RFs of 100–150, 400–500, and 600–750 bp (Fig. 6). However, under prolonged operation, the relative fluorescence intensity of 600–700 bp T-RFs decreased markedly and the 400–500 bp T-RFs became the predominant bacteria community after 80 d, accounting for >60% of the relative abundance of strains. The Shannon–Weaver diversity (H), richness (S) and evenness (E) indices were calculated using data generated from T-RFLP fingerprints (Table 3). The result indicated that the bacterial community under long-term operation with complete sludge retention had fewer species and uneven distributions. This phenomenon may be explained by the fact that the decreasing F/M ratios with time probably led the competition among the microorganisms and therefore resulted in lower species diversity of the bacterial community [12]. In terms of ammonia-oxidizing bacteria, the 354 T-RF was the dominant fingerprint fragment on 9th and 148th days, accounting for 73.1% and 69.4% of relative abundance, respectively (data not shown). The 491 T-RF contributed 14.8% on 9th day; however, it was not found in the MBR on 148th day. The 334 T-RF (12.0%), not the 491 T-RF, was the second most dominant fragment on 148th day. Although microbial community diversity

Fig. 6. Bacterial T-RLFP profiles of the oxic activated sludge samples in the A1 /A2 /OMBR system during the operating period.

decreased during the test, pollutant removal efficiency and total biological activity were not affected. 3.4. Suspension characteristics The average diameter of microbial flocs in the oxic reactor remained roughly constant (38.2 ± 2.5 ␮m; n = 5), and was slightly smaller than the mean value of 48.5 ± 2.4 ␮m (n = 5) in the control system. The smaller sludge floc size was probably attributed to the intensive turbulence in MBR [37]. The a of oxic sludge samples increased with time (data not shown) and had a weak dependence on sludge concentration (Fig. 7). Fig. 7 showed that a increased less than proportionally with the increase of sludge concentrations, indicating a decreasing influence of higher solid concentrations on the apparent viscosity. The Ostwald model was employed for regression of sludge a and

Table 3 Shannon–Weaver diversity (H), richness (S) and evenness (E) values of the oxic activated sludge samples in the A1 /A2 /O-MBR system during the operating period.

Fig. 5. Variation of MLSS concentration and MLVSS/MLSS ratio with time in the anoxic and oxic reactors of the A1 /A2 /O-MBR system.

Operating day

H

S

E

9th day 80th day 148th day

1.04 0.89 0.89

18 15 14

0.83 0.78 0.75

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Fig. 9. Changes in trans-membrane pressure (TMP) over operation time.

Fig. 7. Variation of the apparent viscosity with sludge concentration in the oxic reactor of the A1 /A2 /O-MBR system (the shear rate: 16.7 s−1 ).

MLSS concentration data [32] as follows:

 dv (−8.08 MLSS−1.63 )

a = (1 + 3.64 MLSS0.17 )

dy

for the control system (Fig. 8). The concentrations of supernatant DOC in both the MBR and CAS systems were close, whereas those of supernatant COC were significantly higher in the MBR (28–109 mg/L; n = 32) than in the CAS system (18.4 ± 4.8 mg/L; n = 32) (Fig. 8). In other words, the supernatant organic matter was abundant in the MBR suspension, primarily contributed by the colloidal organic fraction. 3.5. Membrane performance

(1)

The corresponding residual sum of squares (least square method, MATLAB 7.0TM ) was 0.87, indicating a satisfactory correlation between a and MLSS (Fig. 7). The sludge a was significantly lower than that identified by Pollice et al. [32] (Fig. 7). The low content of filamentous bacteria in the sludge may account for this observation. A low-sludge a could facilitate circulation of mixed liquor and shear stress at the filtration cake surface [38], thereby yielding a low-filtration resistance [39]. The supernatant TOC contents in the oxic reactor suspension fluctuated at 70–224 mg/L, and were always higher than those

During the entire 160-d operation, membrane flux was maintained at 6.75 L/(m2 h) with a constant aeration intensity of 2 m3 /h. Cleaning was conducted once when trans-membrane pressure (TMP) reached 0.02 MPa (Fig. 9). On 62nd and 107th days, the permeabilities of the fouled membrane were only 3.6% and 4.5% that of clean membrane, respectively; hence, ex situ physical cleaning was conducted, followed by chemical cleaning. The corresponding permeabilities after physical cleaning were restored to 44.0% and 45.0%, and, after subsequent chemical cleaning, to 92.7% and 93.2%, respectively. That is, the TMP increase was caused by both the physically reversible cake layer (40–41%) and irreversible gel layer (48–49%). On 123rd and 127th days, only physical cleaning was performed to restore membrane permeability when the TMP reached 0.0025 MPa, which was significantly lower than the previous cleaning criterion (TMP > 0.02 MPa). The TMP was appropriately restored (Fig. 9). In other words, the initial deposit layer on the membrane was easily removed by physical cleaning. Further filtration may force fine particles to deposit into the membrane or induce intensive precipitation of dissolved organic matter within the membrane that requires chemical cleaning. The average increase rates of TMP (kTMP ) at the initial stage (first 15 d) of each run were utilized as indicators of membrane filterability. The kTMP values increased from 115 Pa/d in the first run to 140–160 Pa/d in the second run, indicating a stable membrane fouling behavior for long-term operation. Pearson correlation analyses were performed to identify correlations between mixed liquor properties (COC, DOC, TOC, MLSS, MLVSS/MLSS, a ) and kTMP . Among all properties involved, supernatant COC most strongly corTable 4 Results of pearson correlation analyses between the suspension properties and membrane filterability (kTMP ).

Fig. 8. Variation of oxic supernatant TOC, DOC, and COC with time.

Properties

R

P

COC DOC TOC MLSS MLSS/MLVSS a

0.956 0.798 0.911 0.729 0.562 0.841

0.0098 0.1054 0.0312 0.1619 0.3246 0.0739

Note: average values of various properties of individual runs during the first 15 d were used for Pearson correlation analysis.

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Fig. 10. Variation of relative flux during filtration of MBR supernatant fractions (TOC = 53.7 mg/L, DOC = 24.5 mg/L, COC = 29.2 mg/L, T = 25 ◦ C, P = 5 kPa).

related with kTMP (P = 0.0098; highest absolute value of R = 0.956) (Table 4). This analysis indicated that the colloidal organic fraction of supernatant contributed markedly to membrane clogging. Independent dead-end filtration tests using MBR supernatant (containing colloidal and dissolved organic fractions) revealed significant flux decline over time (Fig. 10). However, the supernatant with the removal of colloids yielded a negligible flux decline. This result implied a high-membrane-fouling potential of the colloidal organic fractions in supernatant, which was consistent with the findings in the above section. On the other hand, even though the fouling potential of the dissolved organic fractions was not obvious in the dead-end filtration tests, the contribution of these fractions to membrane fouling could not be neglected [29], since their fouling effect may be exerted under long-term operation in MBR system. 4. Conclusions The proposed A1 /A2 /O-MBR system operated with complete sludge retention was capable of stably removing 87.9 ± 1.6% of COD, 99.4 ± 0.3% of turbidity, and 99.7 ± 3.5% of NH4 + -N from highly toxic coke wastewater. The anaerobic reactor (A1 ) primarily converted refractory and inhibitory compounds into biodegradable organic substances. The anoxic reactor (A2 ) denitrified nitrate and removed >50% of COD; the oxic reactor degraded 33% of COD and nitrified ammonium-nitrogen to nitrate. The membrane rejected 7.1% of COD and most SS, creating to a low food-to-microorganism environment to degrade refractory substances and reduce sludge reduction. Microbial diversity decreased over time; however, neither pollutant removal efficiency nor total biological activity was adversely affected. Membrane fouling occurred during MBR operation was mostly attributable to colloidal fractions of supernatant in suspension. The initial deposit of particles can be removed easily by physical cleaning. Prolonged operation yielded irreversible clogging, which can be rectified by chemical cleaning. Acknowledgement This work was supported by the National Science Fund for Distinguished Young Scholars (no. 50725827). References [1] T.D. Glen, E.R. Bruce, A. Samer, A. Gianni, Are membrane bioreactors ready for widespread application? Environ. Sci. Technol. 39 (2005) 385A–408A. [2] S. Judd, The status of membrane bioreactor technology, Trends Biotechnol. 26 (2008) 109–116.

63

[3] W.B. Yang, N. Cicek, J. Ilg, State of the art of membrane bioreactors: worldwide research and commercial applications in North America, J. Membr. Sci. 270 (2006) 201–211. [4] C. Visvanathan, R. Ben Aim, K. Parameshwaran, Membrane separation bioreactors for wastewater treatment, Crit. Rev. Environ. Sci. Technol. 30 (2000) 1–48. [5] B. Marrot, A. Barrios-Martinez, P. Moulin, N. Roche, Industrial wastewater treatment in a membrane bioreactor: a review, Environ. Prog. 23 (2004) 59–68. [6] K. Brindle, T. Stephenson, The application of membrane biological reactors for the treatment of wastewaters, Biotechnol. Bioeng. 49 (1996) 601–610. [7] C.H. Xing, W.Z. Wu, Y. Qian, E. Tardieu, Excess sludge production in membrane bioreactors: a theoretical investigation, J. Environ. Eng. ASCE 129 (2003) 291–297. [8] R. Liu, X. Huang, J.Y. Xi, Y. Qian, Microbial behaviour in a membrane bioreactor with complete sludge retention, Process Biochem. 40 (2005) 3165– 3170. [9] A. Pollice, G. Laera, D. Saturno, C. Giordano, Effects of sludge retention time on the performance of a membrane bioreactor treating municipal sewage, J. Membr. Sci. 317 (2008) 65–70. [10] H.Y. Li, M. Yang, Y. Zhang, T. Yu, Y. Kamagata, Nitrification performance and microbial community dynamics in a submerged membrane bioreactor with complete sludge retention, J. Biotechnol. 123 (2006) 60–70. [11] S. Rosenberger, U. Kruger, R. Witzig, W. Manz, U. Szewzyk, M. Kraume, Performance of a bioreactor with submerged membranes for aerobic treatment of municipal wastewater, Water Res. 36 (2002) 413–420. [12] T.W. Tan, H.Y. Ng, S.L. Ong, Effect of mean cell residence time on the performance and microbial diversity of pre-denitrification submerged membrane bioreactors, Chemosphere 70 (2008) 387–396. [13] M.K. Ghose, Complete physico-chemical treatment for coke plant effluents, Water Res. 36 (2002) 1127–1134. [14] E. Maranon, I. Vazquez, J. Rodrıguez, L. Castrillon, Y. Fernandez, H. Lopez, Treatment of coke wastewater in a sequential batch reactor (SBR) at pilot plant scale, Bioresour. Technol. 99 (2008) 4192–4198. [15] M. Zhang, J.H. Tay, Y. Qian, X.S. Gu, Coke plant wastewater treatment by fixed biofilm system for COD and NH3 -N removal, Water Res. 32 (1998) 519– 527. [16] Y.M. Li, G.W. Gu, J.F. Zhao, H.Q. Yu, Y.L. Qiu, Y.Z. Peng, Treatment of coke-plant wastewater by biofilm systems for removal of organic compounds and nitrogen, Chemosphere 52 (2003) 997–1005. [17] D.H. Park, D.S. Lee, Y.M. Kim, J.M. Park, Bioaugmentation of cyanide-degrading microorganisms in a full-scale cokes wastewater treatment facility, Bioresour. Technol. 99 (2008) 2092–2096. [18] J.L. Wang, X.C. Quan, L.B. Wu, Y. Qian, H. Werner, Bioaugmentation as a tool to enhance the removal of refractory compound in coke plant wastewater, Process Biochem. 38 (2003) 777–781. [19] R. Qi, K. Yang, Z.X. Yu, Treatment of coke plant wastewater by SND fixed biofilm hybrid system, J. Environ. Sci. 19 (2007) 153–159. [20] Y.M. Kim, D.H. Park, C.O. Jeon, D.S. Lee, J.M. Park, Effect of HRT on the biological pre-denitrification process for the simultaneous removal of toxic pollutants from cokes wastewater, Bioresour. Technol. 99 (2008) 8824–8832. [21] P. Lai, H.Z. Zhao, Z.F. Ye, J.R. Ni, Assessing the effectiveness of treating coking effluents using anaerobic and aerobic biofilms, Process Biochem. 43 (2008) 229–237. [22] R.W. Walters, R.G. Luthy, Liquid/suspended solid phase partitioning of polycyclic aromatic hydrocarbons in coal coking wastewaters, Water Res. 18 (1984) 795–809. [23] K. Yamamoto, M. Hiasa, T. Mahmood, T. Matsuo, Direct solid–liquid separation using hollow fiber membrane in an activated sludge aeration tank, Water Sci. Technol. 21 (1989) 43–54. [24] E.B. Muller, A.H. Stouthamer, H.W. Verseveld, D.H. Eikelboom, Aerobic domestic wastewater treatment in a pilot plant with complete sludge retention by crossflow filtration, Water Res. 299 (1995) 1179–1189. [25] P.C. Sridang, A. Pottier, C. Wisniewski, A. Grasmick, Performance and microbial surveying in submerged membrane bioreactor for seafood processing wastewater treatment, J. Membr. Sci. 317 (2008) 43–49. [26] APHA, AWWA, WEF, Standard Methods for the Examination of Water and Wastewater, 20th ed., APHA, Washington, DC, USA, 1998. [27] S.G. Joanna, G. Krist, D. Carl, V. Peter, V. Willy, Nitrification monitoring in activated sludge by oxygen uptake rate (OUR) measurements, Water Res. 30 (1996) 1228–1236. [28] H.D. Park, R.N. Daniel, Evaluating the effect of dissolved oxygen on ammoniaoxidizing bacterial communities in activated sludge, Water Res. 38 (2004) 3275–3286. [29] E. Bouhabila, R. Ben Aim, H. Buisson, Fouling characterisation in membrane bioreactors, Sep. Purif. Technol. 22 (2001) 123–132. [30] X.Y. Li, H.P. Chu, Membrane bioreactor for the drinking water treatment of polluted surface water supplies, Water Res. 37 (2003) 4781–4791. [31] M.W. Lee, Y.J. Park, Biological nitrogen removal from coke plant wastewater with external carbon addition, Water Environ. Res. 70 (1998) 1090– 1095. [32] A. Pollice, C. Giordano, G. Laera, D. Saturno, G. Mininni, Physical characteristics of the sludge in a complete retention membrane bioreactor, Water Res. 41 (2007) 1832–1840. [33] A. Pollice, G. Laera, M. Blonda, Biomass growth and activity in a membrane bioreactor with complete sludge retention, Water Res. 38 (2004) 1799– 1808.

64

W.-T. Zhao et al. / Journal of Membrane Science 330 (2009) 57–64

[34] R. Witzig, W. Manz, S. Rosenberger, U. Kruger, M. Kraume, U. Szewzyk, Microbiological aspects of a bioreactor with submerged membranes for aerobic treatment of municipal wastewater, Water Res. 36 (2002) 394–402. [35] N. Tixier, G. Guibaud, M. Baudu, Determination of some rheological parameters for the characterization of activated sludge, Bioresour. Technol. 90 (2003) 215–220. [36] F.G. Meng, H.M. Zhang, F.L. Yang, Y.S. Li, J.N. Xiao, X.W. Zhang, Effect of filamentous bacteria on membrane fouling in submerged membrane bioreactor, J. Membr. Sci. 272 (2006) 161–168.

[37] T.H. Bae, T.M. Tak, Interpretation of fouling characteristics of ultrafiltration membranes during the filtration of membrane bioreactor mixed liquor, J. Membr. Sci. 264 (2005) 151–160. [38] G. Laera, C. Giordano, A. Pollice, D. Saturno, G. Mininni, Membrane bioreactor sludge rheology at different solid retention times, Water Res. 41 (2007) 4197–4203. [39] T. Itonaga, K. Kimura, Y. Watanabe, Influence of suspension viscosity and colloidal particles on permeability of membrane used in membrane bioreactor (MBR), Water Sci. Technol. 50 (2004) 301–309.