Development of salt-tolerant microbial consortium during the treatment of saline bisphenol A-containing wastewater: Removal mechanisms and microbial characterization

Journal of Water Process Engineering 32 (2019) 100949

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Development of salt-tolerant microbial consortium during the treatment of saline bisphenol A-containing wastewater: Removal mechanisms and microbial characterization

T

Masoumeh Golshana,b, Sahand Jorfia,b, Neamatollah Jaafarzadeh Haghighifarda,b, ⁎ Afshin Takdastana,b, Shokouh Ghafaric, Soodabeh Rostamid, Mehdi Ahmadia,b, a

Environmental Technologies Research Center, Ahvaz Jundishapur University of Medical Sciences, Ahvaz, Iran Department of Environmental Health Engineering, Ahvaz Jundishapur University of Medical Sciences, Ahvaz, Iran Infectious Diseases Research Center, Birjand University of Medical Sciences, Birjand, Iran d Nosocomial Infection Research Center, Isfahan University of Medical Sciences, Isfahan, Iran b c

A R T I C LE I N FO

A B S T R A C T

Keywords: Bisphenol A Biokinetic parameters Sequencing batch reactor (SBR) Saline wastewater treatment Microbial community

Recalcitrant compounds and high salinity in industrial wastewaters are two major inhibitory parameter against the bacterial metabolism leading to necessity for the application of halotolerant microorganisms in biological treatment technologies. Hence, this study focused on the biological treatment of saline bisphenol A (BPA)containing wastewater at different total dissolved solids (TDS) (5, 10 and 15 g/L) with BPA concentration of 50 mg/L. Three sequencing batch reactors (SBR) were operated which applied different experimental conditions during 9 months. The pure adsorption behavior of BPA onto sludge was described via Freundlich isotherm model in batch experiments. The operation of sequencing reactors with 5 days hydraulic retention time (HRT) indicated that optimal removal rates exceeded 96.3%, 88% and 57% for BPA, COD, TOC, respectively, which was attained at high salinity. The organisms responsible for BPA removal appeared to be more sensitive to different operating conditions than changes in salinity; the BPA removal efficiency decreased from 96.3% to 69.8%, when HRT decreased from 5 days to 0.25 days. Results indicated that biodegradation was the predominant process for BPA removal. The biokinetic parameters in saline substrates were in the range of Y = 0.54-0.61 (mg VSS/mg BPA), kd = 0.006-0.013 (1/d), Ks = 8.94–13.6 (mg BPA/L) and μm = 0.3-0.4 (1/d). Identification of mixed consortium at high salinity was included the species Pseudomonas aeruginosa, three different Serratia marcescens, Bordetella muralis, and Bacillus subtilis by using 16S rRNA-analysis.

1. Introduction The increasingly water demand, stringent discharge standards and constraints of water resources have caused an interest to discover the innovative wastewater treatment technologies. The most industrial concern is by far the saline wastewaters treatment rich in industries such as petroleum, petrochemical, pharmaceutical, landfill leachate, pickling, and agricultural pesticides and herbicides [1]. Saline wastewater has been unveiled by considerable quantities of inorganic salts and high values of xenobiotic organic compounds, which are usually non-biodegradable and resistant against common biological treatment processes inhibited by salt. Thus, they rose a great concern about their contribution to adverse effect on almost all environmental supplies and application of halotolerant bacteria are required in alternative

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biological processes to enhance biodegradation rate of target recalcitrant organics in saline environment [2,3]. Biodegradation is an efficient technique capable of reducing the overall toxicity of organic substances in saline environment. However, salinity (> 1% salt) gives rise to plasmolysis, and/or death of microorganisms, reduces degradation kinetics, inhibits the nitrification and forces the operation at low activated sludge loadings [4,5]. Saline environment involves in decantation problems, caused by higher density [2]. Nowadays, researchers are trying to identify novel bacterial isolates from saline environment to prepare halotolerant consortiums as one of the most interesting approaches for improving the performance of activated sludge systems [6]. Besides, gradual biomass acclimation to high salinity have led to the obtention of satisfactory quality of effluent [7]. Effect of salinity on variations bacterial density as well as removal

Corresponding author at: Environmental Technologies Research Center (ETRC), Ahvaz Jundishapur University of Medical Sciences (AJUMS), Ahvaz, Iran. E-mail address: [email protected] (M. Ahmadi).

https://doi.org/10.1016/j.jwpe.2019.100949 Received 5 July 2019; Received in revised form 29 August 2019; Accepted 8 September 2019 2214-7144/ © 2019 Elsevier Ltd. All rights reserved.

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and destruction of organic pollutants has been considered in the literature [8]. It was reported that acclimation to high salinity raised the halotolerant and/or halophilic bacteria, resulting in enhancement of the bioreactor performance [3]. One of the most common organic compounds in saline environment generated in petrochemical plants is Bisphenol A (BPA) [9,10] originating from condensation of major constituent phenol and acetone under low pH and high temperature conditions and is listed as a xenobiotic leads to hormonal imbalance in cells with the highest global volume consumption by 2022 [11]. The point that should be considered is that BPA is required as an intermediate to produce rubber and plastics, glue, ink, poly(vinyl chloride), unsaturated polyesters, and polyacrylate resins. The environment receives BPA via effluents of industrial and municipal wastewater treatment plants (WWTP), waste plastics in landfill leachate, activated sludge processing units, sediments of river or seawater [12]. This leads to the ubiquitous occurrence of BPA in the environment making it a major focus of attention. In activated sludge treatment processes, adsorption to biomass accounts for a considerable removal of endocrine-disrupting compounds [8], leading to major concerns for sludge management [13]. Studies indicated that BPA tends to sorption to the biomass originated from its hydrophobicity (log Kow 4) [14–16]. While, some research on biodegradation of BPA have focused only on isolation and characterization of bacterial isolates in saline environment [17,18], based on literature, none of them has characterized BPA biodegradation mechanism in such conditions. Eio et al. [19] showed that mixed consortium from activated sludge achieved complete degradation of BPA in batch assays. The wide range of catabolic pathways of mixed consortium provided efficient degradation of BPA and its biodegradation intermediates led to converting BPA to metabolites with much less estrogenic activity. Cydzik-Kwiatkowska et al. [11] investigated the sludge granulation with the potential of removal of BPA-rich-wastewater containing 12 mg/L BPA in sequencing reactors. Former studies were based on application of non-halophilic species, while detailed information on BPA degradation by halophiles and phylogenetic analysis is limited. The lack of detailed information is despite the wide industrial application of BPA that was closely interrelated with saline nature of the effluents. The sequencing batch reactor (SBR) has continued to be a robust process in high saline wastewaters treatment [2,7]. This study was designed to investigate the overall performance and influence of effective operational parameters of SBR in biodegradation of BPA in a saline industrial wastewater. The microbial community at highest salinity and mechanisms of BPA removal were also investigated. In addition, biokinetic coefficients of halotolerant consortium were evaluated.

Fig. 1. Schematic of SBR. Table 1 Operational conditions of SBR operation parameters. Experimental run

Acclimation to

1 2 3 4 5 6

Operation (days) 5 g/L TDS 10 g/L TDS 15 g/L TDS – – – – – –

Flow rate (L/day)

HRT (days)

OLR (g BPA/m3.d)

2 3.3 5 10 20 40

5 3 2 1 0.5 0.25

10 16.7 25 50 100 200

59 112 176 13 17 14 18 15 16

defined as conditions with less than 5–7% variations in effluent characteristics. 2.3. Inoculum preparation For the seed sludge, aerobic activated sludge was obtained from WWTP of Ahvaz, Iran. The salt concentrations of the raw wastewater were adjusted in terms of TDS parameter by adding a desired amount of NaCl purchased commercially. The first reactor received the raw wastewater containing 5 g/L TDS, whereas the other two reactors treated the raw wastewater containing 10 and 15 g/L TDS. The biomass was first incubated with the synthetic substrate containing glucose (400 mg/ L). The desired amounts of NH4Cl and KH2PO4 were added to the tap water in keeping with C:N:P ratio of 100:5:1. Salt build-up was administered by stepwise increase in wastewater salt concentration to the substrate over a time period. The acclimation to change salinity was occurred when the COD removal reached to over 80%; after that the salt concentration was increased further. The periods of acclimation to different TDS concentrations are presented in Table 1. Once the COD removal reached to over 80%, BPA (Sigma–Aldrich > purity 99%) increased stepwise in the raw wastewater to reach concentration of 50 mg/L of BPA during one month. BPA stock solution (1 M) was dissolved in NaOH aqueous media [19]. Sampling was carried out on raw wastewater, effluent and mixed liquor of each SBR. The pH was set in neutral range (7.5 ± 0.2) with NaHCO3.

2. Methods 2.1. Reactor set up Three plexi-glass SBR systems with an operating volume of 10 L and size of 28 × 18 × 19 cm (L × W × H), were operated (Fig. 1). A vacuum pump was employed to inject the synthesized wastewater into the reactors at required flow rate range of 2–40 L/d controlled by programmable logic controllers. The air pumps (Hailea, model ACO-208) with an air flow rate of 6 L/min were used to create turbulent flow in the reactor contents and supply enough dissolved oxygen (DO 3–4 mg/ L) through three stone air diffusers placed at the bottom of the reactor. 2.2. Reactor operation

2.4. Adsorption test protocol To test the effect of operational conditions on BPA removal, various hydraulic retention time (HRT) values of 5 to 0.25 days were considered. The operating conditions during the study are included in Table 1. Feeding, settling and decanting were operated at 15, 30 and 15 min, respectively, during all runs. The steady state conditions were

The batch adsorption experiments were studied in 250-mL conical flasks with inactivated sludge originated from the SBRs. These SBRs were operated at various salinities, devoid of BPA as described above. First, the biomass was centrifuged at 4000 rpm during 10 min followed 2

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spectrophotometer model DR/6000 making proper dilutions and centrifuge (5000 g) minimizing the impact of salinity on COD analysis. TDS, DO and pH were monitored by specific electrodes (Hach Company, USA). BPA concentration was measured by a HPLC instrument (model KNAUER). Before beginning analysis, 50 mL mixed liquor originated from each reactor was centrifuged at 4000 rpm during 10 min for separation of the liquid and solid phases. Liqud samples were filtered with 0.45 μm glass-fiber syringe filters (Whatman). Filtration did not affect the BPA concentration. Subsequently, 100 μL filtrate was analyzed by the HPLC system equipped with ultimate variable wavelength UV detector (2500) connected to a column (100-5 C18; 4.6 mm × 250 mm, 5 μm) in an oven temperature of 30 °C. The mobile phase composed of acetonitrile/Miliporewater (50:50, v/v) at a flow rate of 1 mL/min and the detection wavelength was 214 nm in a retention time of 7.5 min. For determination of adsorbed BPA on the sludge phase, solid pellet was ultrasonicated for 30 min using 20 mL of methanol-dichloromethane (1:1, v:v). Afterwardes, the liquid sample was rotary evaporated at 30 °C to remove the solvent and dissolved again in 2 mL methanol and then analyzed, according to the aforementioned HPLC protocol [23]. The detection limit of BPA was 0.1 mg/L. The intermediates of BPA biodegradation were discerned using LC–MS (Agilent QQQ-6410) analysis.

by decanting the supernatant and rinsing the pellet with mili-Q water thrice. Mercury (II) sulfate (100 mg/L) was used to inactivate the biomass [20]. Afterwards, the required MLSS level was adjusted with tap water and the prepared samples were stored at 4 °C. For the equilibrium time determination, 100 mL sterilized sludge solution with MLSS concentration 2000 mg/L were amended with 50 mg/L of BPA concentration and shaken (IKM 4000 ci, Germany) at 125 rpm and 20 °C. pH was kept at 7.5 ± 0.2. Supernatant samples for measuring BPA concentrations were collected at 5, 15, 30, 60, 90, 120, 300, 600, and 840 min. The concentrations of 5, 10, 20, 30, 40, and 50 mg/L of BPA were considered to develop the sorption isotherms. A blank sample without sludge was considered under the same conditions. Freundlich equilibrium model was applied for fitting the experimental data as expressed in (Eq. (1)): Logqe = LogKF + (1/n) LogCe

(1)

where, Ce (mg/L) and qe (mg/g) are the equilibrated concentration of sorbate in water and sorbent phase, respectively. KF (mg/g) (L/mg)1/n and 1/n are capacity and intensity of sorption, respectively. Clara et. al [20] expressed that the sorption behavior can be explained in terms of the distribution coefficient, KD (L/g), as the ratio of equilibrated sorbate concentration in two phases (Eq. (2)). KD = (qe/Ce)

(2)

2.7. Isolation of halotolerant bacteria 2.5. Biokinetic coefficients

The isolation and selection of the halotolerant microorganisms able to degrade BPA were conducted in consecutive batch assays. The inoculum for the first batch was taken from the biomass of SBR operated at 15 g/L TDS and then incubated into 100 mL culture medium in a 250mL flask. The culture medium was composed of K2HPO4, 6.3; KH2PO4, 1.8; NH4Cl, 1; MgSO4.7H2O, 0.1; CaCl2.H2O, 0.1; FeSO4.7H2O, 0.1; MnSO4.H2O, 0.1 (in g/L) and 1 ml/L of micronutrient solution, pH 7.0. The composition of micronutrient was H3BO3, 0.03; ZnSO4.7H2O, 0.01; CoCl2.6H2O, 0.02; Na2MoO4, 0.006; CuSO4.2H2O, 0.001 (in g/L). NaCl was added to the culture to obtain a concentration of 15 g/L TDS and then sterilized. Experimental flasks were adjusted to 100 mg/L BPA and then maintained at 31 °C in the dark with a reciprocal shaker during one week. BPA was sterilized (Millipore membrane with 0.2-μm pore size) for removing bacterial. The bacterial growth was detected by the variation of optical density at 600 nm. After one week of cultivation, 5 mL enriched sample was re-incubated to another flask with 95 mL sterile medium under the same conditions [24]. Then, for the next batch assay, the inoculum was obtained from previous one. This step was performed consecutively four times and then the culture was streaked on mineral agar plates, containing 100 mg/L BPA and 15 g/L NaCl based on TDS, and incubated at 31 °C for 3 days. Large distinct colonies were isolated in pure culture on plates.

The kinetics model used in this study is based on Monod kinetics relationship (Eq. (3)) under steady state conditions (dS/dt = 0, dX/ dt = 0) which the substrate utilizing and biomass change rate are obtained by Eqs. (4) and (5), respectively: μ = μm [S/Ks + S]

(3)

- dS/dt = [S0 – S / θH] – [μX/Y]

(4)

dX/dt = [(- 1/θc) - μ - kd]X → {μ = (1/θc) + kd → (1/θc) + kd = μm [S/Ks + S]} (5) By combining Eqs. (4) and (5), the substrate utilization rate can be estimated as seen in Eqs. (6) and (7). (S0 – S)/ θHX = [1/Yθc] + [kd/Y] = U

(6)

θc/1+ θckd = [Ks/μm](1/S) + 1/μm = 1/U

(7)

Accordingly, the values of kinetic coefficients Y (sludge yield, mgVSS/mgBPA), kd (decay coefficient, 1/d), km (substrate utilization rate, mgBPA/mgVSS.d), Ks (constant affinity for substrate, mgBPA/L) and μm (maximum growth rate, 1/d) over TDS concentrations can be determined by linear regression analysis (Eqs. (6) and (7)). As a result, plot of values of 1/θc (θc = VX/QXe) indicates the ratio of biomass concentration in the reactor to biomass concentration exiting the system per day) aganist the specific substrate utilization rate (U) gives Y and kd values via the slope and the y-intercept of the line of best fit to plotted data. Moreover, the plot of 1/U against 1/S (S is substrate concentration in the treated effluent, mg/L) determines slope Ks and intercept μm. X signifies biomass concentration in the reactor (mg VSS/ L) and HRT (θH = V/Q) is expressed by division of the working volume of the bioreactor and the raw wastewater flow rate [21].

2.8. The study of pure bacterial isolates The identification of bacteria was performed according to culturalmorphological and biochemical features using Bergey's Manual of Systematic Bacteriology [25] as well as 16S rRNA gene sequencing. DNA was exteracted using phenol-chloroform extraction. The 16S rRNA genes were amplified with universal primers fD1 (5′-AGAGTTTGATCC TGGCTCAG-3′) and rD1 (5′- AAGGAGGTGATCCAGCC-3′). The sequences were derived from PCR amplification as: 95 °C for 5 min, 35 cycles at 95 °C for 30 s (denaturing step), 55 °C for 30 s (annealing) and 72 °C for 1.5 min (extension), and a final extension at 72 °C for 15 min. The PCR substances were purified and sequenced by using a 3730XL DNA analyzer instrument (Applied Biosystems) under contract by bioneer Inc (South Korea). All fragmented sequences were edited and assembled using DNA Sequence Assembler v4 (2013). Sequence analysis was carried out by using the EzTaxon server (http://www.ezbiocloud. net/eztaxon) [26]. Phylogenetic analyses were accomplished in

2.6. Analytical methods Total organic carbon (TOC) was assigned using a TOC-VCHS/CSN analyser (Shimadzu, Japan). Chemical oxygen demand (COD) by colorimetric method (5220-D), mixed liquor suspended solids (MLSS) and mixed liquor volatile suspended solids (MLVSS) were tested following Standard Methods [22]. COD was measured through a Hach 3

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Fig. 2. Adsorption isotherm model of BPA determined with sludge at 15 g/L TDS concentration.

Molecular Evolutionary Genetics Analysis (MEGA6) software package by maximum likelihood algorithm using Kimura-2-parameter distances [27] and 1000-bootstrap replication.

3. Result and disscussion 3.1. Batch experiments Due to low volatility of BPA (vapour pressure 5.34 × 10−7 mmHg, at 25 °C and Henry’s constant 10 μPa m3 mol-1) [28], abiotic removal owing to volatilization could be ignored. No BPA removal was observed in biomass free experiments. In the first 30 min, the BPA adsorption on the chemically inactivated sludge was so quick (data not shown) and the BPA removal of 65% was obtained, while there was no significant change after 1 h in all experiments. Determination of adsorption isotherm was carried out during 1 h contact time. The adsorption isotherm of sludge at 15 g/L TDS concentration can be observed in Fig. 2. It was clear that the adsorption of BPA to sludge followed Freundlich model. The sorption data charecterizing both Freundlich and linear equations for three salinities and the inactivated sludge are presented in Table 2. The values of Freundlich constant log Kf recorded at different salinities varied from -0.771 to -0.615 for 5 to 15 g/L TDS (Table 2), respectively which were quite similar to those of Clara et al. [20] (-0.614) and Seyhi et al. [29] (-0.637) were obtained. Our results pointed out that the distribution coefficients log KD on different sludges were 1.97–2.42, 2.03–2.42 and 1.97–2.27 for 5, 10 and 15 g/L TDS, respectively (Table 2). This parameter has the concentration-dependent behavior, which shows a decreasing trend with increasing free BPA concentrations [20]. Log KD was on a par with the value (2.52) yielded at BPA concentrations in low range (2–100 μg/L) [13]. Chen et al. [16] indicated that log KD values of BPA ranged from 2 to 2.75, applying high concentration range (2.5–40 mg/L). Urase and Kikuta [30] stated that log KD values of BPA on activated sludge were between 2.33 and 2.8. In comparison with the reported values in literatures, our values of KD were consistent with other values. From these results, it can be postulated that the sludge had a high adsorption potential for this compound without being saturated in all tested concentrations [20].

Fig. 3. BPA removal (%) in SBR systems with different salinity concentration: (a) 5 g/L TDS, (b) 10 g/L TDS and (c) 15 g/L TDS.

Our results suggested that the salinity had no influence on the adsorption effeciency of BPA (Table 2). To the extent of our knowledge, to date, this is the first report about comparison between sorption constants of BPA at different salinity levels. The impact of high salinity on adsorption of BPA in an aerobic sludge was negligible [8]. Song et al. [31] observed a fluctuated rise in the BPA concentration in a municiple digester sludge during anaerobic membrane bioreactor treatment as salinity increased (0–15 gNaCl/L). These studies contributed to different and contradictory conclusions.

Table 2 Results of Freundlich equilibrium parameters and distribution coefficient for BPA with different TDS concentrations. Target Compound

BPA

TDS (g/L)

5 10 15

Freundlich equation

Log KD

KF (mg/g) (L/mg)1/n

1/n

R2

0.204 0.27 0.242

0.771 0.702 0.665

0.99 0.988 0.998

3.2. Removal of Bisphenol-A during aerobic SBR treatment

1.97-2.42 2.03-2.42 1.97-2.27

3.2.1. Effect of salinity The variations of BPA in SBR systems showed that the majority of 4

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TDS, suggesting that the biomass consortia made an adjustment to BPA biodegradation in saline conditions. The mass balance calculations demonestrated that the fraction of BPA quantified in mixed liquior accounted for 0.3%, in the first experimental phase, while the effluent BPA concentration was 3.9%. Fig. 4 showed the change to lower HRTs partly developed the sorption mechanism, causing the accumulation of BPA in the sludge phase, making it persistent against biodegradation. Concequently, adsorption to the sludge obtained for about 1.9%, when HRT of 6 h was applied and mean removal of 67.94% was occurred in the system operating at salinity up to 15 g/L TDS. This confirms that biological removal was the principle mechanism of BPA removal. Higher HRTs not only developed biodegradation, but also was preferable for the BPA desorption from sludge phase [33]. In other words, BPA biodegradation was facilitated by adsorption phenomenon during SBRs operation with different salinities where the long HRT of the sequencing systems enhanced the BPA bioconversion, due to adsorption to sludge phase [34,35]. In another research, Guerra et al. [15] found that sorption tendency of BPA to sludge was strongly affected by HRT, in which longer HRTs resulted in less sorption, due to improvement of degradation, and the fractions of BPA partitioning to sludge was reduced. Nacheva and Sotelo [36] also reported similar observations, who obtained higher polycyclic aromatic hydrocarbons (PAHs) removal (about 61%) and lower contriobution of PAHs sorption to the sludge, in the MBR operated with long HRT. It seems that high salinity had a negligible impact on BPA adsorption in the sludge phase (Fig. 4), and consequently the adaptation of biomass to salinity levels results in enhancing the process performance in TDS concentrations studied here which is more coincied with the establishment of endured microbial community [3]. This also corroborated the results of adsorption (Table 2). The low accumulation of BPA in the sludge phase when salinity add up to 16.5 g/L NaCl has been observed by Wenhai Luo et al. [8].

BPA was degraded during SBRs operation. High efficiencies of 93, 95 and 96.3% were obtained at salinity values of 5, 10 and 15 g/L TDS, respectively (Fig. 3). This difference could be ascribed to the acclimatization of biomass consortia that was implicated in BPA biodegradation to the saline conditions. Therefore, the increased overall BPA removal efficiency in high salinity was possibly associated with the development of salt-adapted microorganisms that especially attacked to the compounds (Fig. S1). As a result of the accumulation of potassium, glycerol, betaine and amine acids within the cell, the cellular ionic strength is increased, thus enabling the halotolerant bacteria to balance the cellular ionic strength and external environment [3]. In addition, osmoregulation by inducing the ATP formation could stimulate the obligatory energy-dependent reactions in the cells [7]. The obtained results were very close to those of Luo et al. [8] and Zhao et al. [1]. The aforementioned BPA removal was comparable with reported values in previous research about full-scale WWTPs. In a comprehensive survey of 25 Canadian WWTPs, Guerra et al. [15] reported that the removal of BPA was above 70%. 3.2.2. Effect of HRT The enhanced removal of BPA could be acquired as a result of HRT of 5 days and an organic loading 10 g BPA/m3.d for all reactors. Moreover, the BPA removal rate was decreased significantly with reduction of HRT values from 3 to 0.25 days (Fig. 3). Results indicated that HRT was an important factor for BPA removal which were consistent with a rapid decrease in the efficiency of MBR when OLR increased [29]. In contrast, Chen et al. [16] expressed the performance robustness at lower HRTs during the BPA removal in MBR. 3.2.3. Removal mechanism To investigate the removal mechanisms and fate of BPA, the specific BPA content was also determined in the biomass of three SBRs during the experimental period. Using the measured BPA concentration in aquous solution and adsorbed onto the sludge, the mass balance of BPA in the systems was conducted. It was assumed that the biodegradation and sorption to sludge were the main routes governing the rate of BPA removal observed in this study. BPA is relatively non-volatile and has a high tendency towards sorption to the sludge matrix. Using Eq. (8), the value of biodegraded BPA was set-up as the mass balance around the systems [29,32]: Min = Mbio + Mads + Meff

3.3. Kinetic study For effective designing and controlling the full-scale reactors, kinetic analysis is essential to ascertain the biokinetic coefficients using the data of laboratory or pilot-plant studies. The efficiency of SBRs in treating 5, 10 and 15 g/L TDS saline wastewater under varying HRTs in the range of 5-0.25 days corresponding to OLR ranged from 10 to 200 g BPA/m3.d were fitted to Monod model, according to the average steady-state operating conditions (Fig. 5). At high salinity contents, the μmax of BPA-degrading biomass was found to be greater (Table 3). It shows that the inhibitory effect of salinity was fewer on the biological activity during operational conditions perused here, which provided higher growth rate at high salinity that resulted in higher BPA removal rate. This is in agreement with the relevant literature [4,37]. The most simple explanation for higher growth rate in the prolonged exposure to salinity is changing the bacterial metabolism pathways in aerobic conditions and the adjustment and multiplication of salt-tolerant bacteria in halophilic activated sludge population [4,38]. As the biomass (MLVSS) concentration was raised, the effluent BPA content decreased with increasing TDS in the raw wastewater. In this regards, the maximum COD and TOC removal rates were of 88 and 57%, respectively which occured at 15 g/L TDS concentration. For SBRs operating at 10 and 5 g/L TDS wastewaters, the removal efficiencies of COD and TOC were 82.6, 48% and 77.3, 37.8%, respectively at steady state OLR of 10 g BPA/m3.d and HRT of 5 days. It was observed from Table 3 that Km value at highest salt content (15 gTDS/L) was found to be higher (Km = 0.65 mgBPA/VSS.d) than that the Km value at salinity of 10 and 5 gTDS/L which was due to higher growth rate (μm = 0.4 d−1) at 15 gTDS/L sustained higher BPA removal efficiency of 95.74% compared to 94.6% at 5 gTDS/L, respectively. In compasion with a nonsaline activated sludge which was continously spiked with BPA, Ks values for biodegradation of BPA at different salinity contents (8.9–13.6 mgBPA/L) obtained in this study were higher [39]. This indicated that

(8)

where, Min (mg/d) and Meff (mg/d) are the BPA mass of raw wastewater and effluent, respectively and Mads (mg/d) and Mbio (mg/d) are the adsorbed and biodegraded BPA mass, respectively. From Fig. 4, the major part of BPA was biodegraded (about 95.75%) during the HRT of 5 days in the system operating at salinity up to 15 g/L

Fig. 4. BPA mass balance in the SBR systems. 5

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Fig. 6. Phylogenetic tree based on 16S rRNA sequences. Trees were constructed with a neighbor-joining (NJ) method in the MEGA6 software using the Kimura 2-parameter (K2P) model with pairwise-deletion and 1000 bootstrap replicates. Bootstrap values over 70% are shown. Fig. 5. Estimation of kinetic parameters by BPA-degrading population.

3.4. Characterization of salt-tolerant bacteria the halophil microorganisms may possess a lower affinity to the substrate [38], and/or they may require more substrate for survival [7]. These results are also comparable to that of a pure culture chemostat, which obtained the Ks value of 13 mg/L in a synthethic culture with 150 mg/L of BPA [40]. The observed Y values from operated SBRs varied between 0.54 and 0.61 mgVSS/mgBPA (Table 3). By comparison, Lindblom et al. [39] reported the Y value 0.67 gcell/gCOD in nonsaline BPA-degrading activated sludge, meaninig that no significant changes were happened with increasing salinity of the feed wastewater. In this study, Y and Kd values are smaller than those obtained by Hamoda et al. [4] (1.995 gVSS/gTOC and 0.299 1/d) in 10 g/L NaCl saline wastewaster. The bioaugmented saline activated sludge resulted in Y and Kd values 0.54 mgVSS/mgCOD and 0.014 1/d, respectively [24] which were quite similar to this study. As denoted by Y values, carbon in BPA was oxidized by the bacteria population which was partly consumed for cell division. The part that was left would either convert to the other metabolic intermediates or mineralize to CO2. It should be refered that biomass compatibility to saline environment resulted in an effective organic removal performance for BPA degradation. The obtained experimental data provide an evidence for active role of BPA-degrading biomass in the promotion of process.

Six morphologically distinct bacterial strains with the capability of using BPA solely for their metabolism were derived from the reactor with salinity of 15 g/L, based on TDS concentration on mineral agar plates, as described in Section 2.5. The nucleotide sequences with Genbank accession numbers MK045317, MK045328, MK045320, MK045327, MK045329, and MK049903 for G11-G16 were noticed to be the strians Pseudomonas aeruginosa (G12), three diferent Serratia marcescens (G11, G13, G14), Bordetella muralis (G15), and Bacillus subtilis (G16), respectively, by using 16S rRNA gene sequencing (Fig. 6). Also, some morphological and biochemical features of the obtained isolates are listed in Table S1. The genera members known as BPAdegradaers include Sphingomonas, Pseudomonas, Achromobacter, Novosphingobium, Nitrosomonas, Serratia, Bordetella, Alcaligenes, Pandoraea, Klebsiella, Cupriavidus, Streptomyces, and Bacillus [41]. Among them, genera Sphingomonas and Pseudomonas had imprortant roles in BPA biodegradation in seawater samples in media supplemented with 100 ppm BPA [42]. Bacillus sp. strain is more prevalent among the gram-positive BPA degraders [12] as Bacillus pumilus strain BP-22DK efficiently degraded 100 mg/L BPA within 1 day on a mineral salt (50 g/L NaCl) medium supplied with growth factros (e.g. peptone, beef extract and yeast extract), whereas this strain was not able for growing and degrading BPA as a sole substrate [43]. In the current study,

Table 3 Kinetic coefficients for biodegradation of BPA. Treatment system

Substrate

Salinity (g/L TDS)

Y (mgVSS/mgBPA)

Kd (1/d)

μmax (1/d)

Ks (mgBPA/L)

Kmax (mgBPA/mgVSS.d)

Ref

SBR SBR SBR Pure culture chemostat Activated sludge Granular SBR SBR Activated sludge Activated sludge Batch respirometer

BPA BPA BPA BPA + Growth factors BPA + WWTP BPA + Co-substrates Glucose Glucose Petrochemical Wastewater

5 10 15 – – – 10 30 20-37 20

0.56 0.54 0.61 0.39 0.67 0.19 0.414 2 0.54 0.57

0.006 0.018 0.013 0.033 0.05 – 0.083 0.3 0.014 –

0.3 0.31 0.4 0.18 0.34-0.47 0.13 0.205 0.348 0.66 9.95

12.9 8.94 13.6 13.1 6.8-13 – 0.039 – 1315.6 45

0.53 0.57 0.65 0.46 0.5-0.7 0.68 0.496 0.174 1.23 17.45

This study This study This study [40] [39] [11] [21] [4] [24] [38]

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microcosms [47]. In the second proposed pathway, BPA is cleaved into 4-(2-hydroxypropan-2-yl)phenol and HQ. The monooxygenation of the aromatic results in the CeC bond cleavage which was then generated to 4-(2-hydroxypropan-2-yl)phenol and HQ. This pathway has also been derived from Cupriavidus basilensis JF1, Bacillus sp. strain GZB and the bacterial consortium [19,48,49], where the monooxygenas enzymatic reaction activated by cofactors NADPH and FAD involve in metabolizing BPA. Further degradation of HQ leads to the breakage of aromatic ring, consequently generating oxalic acid. 4-isopropenylphenol was also formed, as a consequent of loss of hydrogen atom from carbocationic isopropylphenol. It is expected that breaking C1-body from 4-(2-hydroxypropan-2-yl)phenol will lead to production of HAP. Further oxidation of 4-isopropenylphenol and HAP undergoes cleavage and produces HBAL and HBA rapidly. Although HBAL and HBA were not noticed, they were presumed to be oxidised to oxalic acid which was noticed. We suggested that oxalic acid might be subsequently metabolised to CO2 and cell biomass.

Bordetella muralis identified in biomass sample is affiliated with the Burkholderiales order in Alcaligenaceae family that can tolerate salinity up to 15 g/L TDS observed here. High abundance of Burkholderiales has been confirmed by increased concentrations of BPA in aerobic granular sludge [11]. The abundance of this order has been also found to be dependent on increasing salinity [44]. Reports also indicate that BPAdegrading Bordetella sp. OS17 from soil can degrade BPA (41% removal) [41]. A novel bacterium Achromobacter xylosoxidans belonging to Alcaligenaceae is found out the potential for aerobic degradation of BPA [18]. Bacteria associated with the Serratia genus have been found in a saline WWTP effluent sample [45]. The results of Matsumura et al. [41] also cleared that Serratia sp. HI10 isolated from soil could degrade 59.3% of BPA at concentration 300 mg/L at the second day. Therefore, the existence of active players of BPA biodegradation in pure cultures likely plays a major role in the system, which may promote the smallscale systems by developing of salt-tolerant BPA-degraders to treat real industrial petrochemical wastewaters where salinity and BPA dosages are within the range of 20 g/L TDS and 50 ppm BPA [46].

4. Conclusion 3.5. BPA degradation intermediates The key points from the obtained results are summarised here: The by-products exmained by LCeMS analysis were included 2,2-bis (4-hydroxyphenyl)-1-propanol (m/z = 244), 4-(2-hydroxypropan-2-yl) phenol (m/z = 152), 4-isopropenylphenol (m/z = 137), 4-hydroxyacetophenone (HAP, m/z = 120.9), hydrochinone (HQ, m/ z = 110.4), and oxalic acid (m/z = 89) (Fig. S1). A possible pathway by bacterial metabolism based on the metabolites and recent studies [12,19] are elucidated here (Fig. 7). The first BPA metabolic pathway connected to Sphingomonas sp. strain MV1 metabolism can be interpereted as the methyl hydroxylation of BPA and formation of 2,3-bis(4hydroxyphenyl)-1,2-propanediol (not detected) from 2,2-bis(4-hydroxyphenyl)-l-propanol. As a result of oxidative degradation, 2,3-bis(4hydroxyphenyl)-1,2-propanediol persumably is cleaved to form 4-hydroxybenzoic acid (HBA). Previous studies of BPA degradation showed the accumulation of these metabolites in the cases like activated sludge

• Investigation of BPA adsorption in batch experiments on the saline • • • • •

biomass showed that increasing salinity had no influence on the adsorption process of BPA. The application of SBR systems clarified that the principal amount of BPA (> 93%) would be degraded through the operation in sequencing reactors mainly via biodegradation, reaching an HRT 5 days and salinity ranging from 5 to 15 g/L. Compatibility of the microbial community to changing salinity gave rise to an enhancement in the overall removal performance. A decrease in HRT affected the process performance negatively. The accumulation of BPA in sequencing reactors was determined to be less than 1.9%. The biokinetic parameters in saline substrates were in the range of

Fig. 7. Proposed pathways illustrating the degradation of BPA by halotolerant consortia. 7

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• •

Y = 0.54-0.61 (mg VSS/mg BPA), kd = 0.006-0.013 (1/d), Ks = 8.94–13.6 (mg BPA/L) and μm = 0.3-0.4 (1/d), accounted for the COD removal of 88% at 15 g/L TDS. A number of key players of BPA biodegradation endured higher salinity was isolated which belonged to the Pseudomonas, Serratia, Bordetella and Bacillus genera. Metabolic intermediates were detected as hydroquinone, 4-isopropenylphenol, 4-hydroxyacetophenone, 4-(2-hydroxypropan-2-yl) phenol and oxalic acid.

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