- Email: [email protected]
polymethyl methacrylate membrane for phenol-laden saline wastewater
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Novel external extractive membrane bioreactor (EMBR) using electrospun polydimethylsiloxane/polymethyl methacrylate membrane for phenol-laden saline wastewater ⁎
Long-Fei Rena, Huu Hao Ngob, Cuina Buc, Chenghao Gec, Shou-Qing Nic, Jiahui Shaoa, , Yiliang Hea a b c
School of Environmental Science and Engineering, Shanghai Jiao Tong University, No. 800 Dongchuan Road, Shanghai 200240, Shanghai, PR China Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW 2007, Australia School of Environmental Science and Engineering, Shandong University, No. 26 South Shanda Road, Jinan 250100, Shandong, PR China
H I GH L IG H T S
EMBR was set up with superhydrophobic/organophilic electrospun membrane. • External EMBR achieved the salt rejection, phenol permeation and biodegradation. • External and ammonium were simultaneously removed in external EMBR for detoxication. • Phenol • Microbial community, gene enumeration and EPS release varied with phenol conc.
A R T I C LE I N FO
A B S T R A C T
Keywords: Extractive membrane bioreactor Electrospun membrane Polydimethylsiloxane Phenol removal Bacterial response
Phenol-laden saline wastewaters can adversely affect water, groundwater, soil, organisms and ecosystems. Given that frequently-used biodegradation process is generally inhibited by salinity, this work aims to solve the problem through a novel configuration of external extractive membrane bioreactor (EMBR) for the objective of simultaneous phenol permeation, salt rejection and biodegradation. Contact angles of 160.9 ± 2.2° (water) and 0.0° (phenol) were observed on the electrospun polydimethylsiloxane/polymethyl methacrylate (PDMS/PMMA) membrane, suggesting this superhydrophobic/superorganophilic membrane was suitable for separating phenol from water-soluble salt. Phenol ranging from 14.1 ± 2.7 to 290.7 ± 10.4 mg/L (stages 1 to 8) was continuously permeated and completely biodegraded in external EMBR under a hydraulic retention time (HRT) of 24 h, which corresponded with detoxification performance improving from 6.3% to 70.5%. After phenol exposure of 8 stages, Proteobacteria and Saccharibacteria became the main phyla for microorganisms. Enumeration of functional genes (phe, amoA, narG, nirS) confirmed that phenol was mainly consumed by denitrifiers and other heterotrophs as the sole carbon and energy source via oxidation and ring cleavage. As bacterial responses, these genes’ proliferation was promoted under low phenol concentrations but inhibited under high phenol concentrations. Meanwhile, results of extracellular polymeric substances revealed that protein was the key substance in toxicity resistance, phenol adsorption and transfer.
1. Introduction Phenol-laden saline wastewaters are widely generated in various industries including tanneries, oil refineries and olive oil mills [1]. The coexistence of phenol and salt in wastewater enhances its toxicity and adversely impacts on ecosystem, such as altering aquatic environments and damaging valuable resources [2]. Moreover, effective treatment of
⁎
phenol-laden wastewater becomes more challenging due to the presence of salinity, which is known as serious inhibitory effects on biodegradation for wastewater [3]. Compared with traditional biodegradation process, extractive membrane bioreactor (EMBR) is a novel combined process of membrane separation and biodegradation for wastewater treatment [4]. A hydrophobic-organophilic membrane was immersed in EMBR to divide
Corresponding author. E-mail address: [email protected] (J. Shao).
https://doi.org/10.1016/j.cej.2019.123179 Received 9 August 2019; Received in revised form 11 October 2019; Accepted 14 October 2019 1385-8947/ © 2019 Elsevier B.V. All rights reserved.
Please cite this article as: Long-Fei Ren, et al., Chemical Engineering Journal, https://doi.org/10.1016/j.cej.2019.123179
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Therefore, a novel external EMBR was firstly set up using an electrospun PDMS/PMMA membrane to treat phenol-laden saline wastewater. Then, the effects of operation conditions (hydraulic retention time (HRT) and concentration) were investigated to achieve efficient phenol separation and biodegradation. The objectives of this study were to: (i) investigate the phenol permeation and salt rejection; (ii) examine the pollutant removal and detoxification; (iii) evaluate the bacterial responses (microbial community, functional gene and extracellular polymeric substances (EPS)) under different phenol-laden saline wastewaters and (iv) elucidate the degradation pathway of phenol. This study should provide better understanding of external EMBR to facilitate phenol-laden saline wastewater treatment.
it into wastewater and bioreactor. Organic pollutants in wastewater could transport through the selectively permeable membrane into bioreactor via solution-diffusion for subsequent biodegradation [4,5]. Meanwhile, inorganic pollutants (e.g., salinity, acid, alkali) are separated by the hydrophobic membrane surface, which can probably affect biodegradation efficiency [4,5]. Therefore, bioreactor could be adjusted to suitable condition for the internal microorganisms to accelerate the removal of permeated organic pollutants and provided nutrients (e.g., nitrogen, phosphorus). These attributes make EMBR suitable for the treatment of toxic recalcitrant organic wastewater where harsh conditions exist (e.g., phenol-laden saline wastewater). EMBR has been widely reported in the treatments of phenol saline wastewaters [4–6]. All these attempts on EMBR used the internal-installed membrane configuration, which facilitated the simultaneous phenol separation and phenol biodegradation. Though the units of wastewater, selectively permeable membrane and microorganisms are excessively integrated in this configuration, it is difficulty to control phenol separation or phenol biodegradation precisely. Consequently, a minor problem in wastewater, selectively permeable membrane or microorganisms may collapse the whole EMBR system. Moreover, the internal-installed membrane configuration is prone to make membranes being subjected to biomass adhesion even biofouling from the microorganisms in bioreactor [7–9]. In comparison, the external-installed membrane configuration could address the above issues effectively: 1) relatively independent units of wastewater, selectively permeable membrane and microorganisms are easier to control, 2) adhesion of biomass or colloidal organic matter to the membrane surface would be completely avoided to free membrane from biofouling [10]. Though the external-installed membrane configuration has been successfully used in MBR fields [10,11], this membrane configuration has not been used in EMBR to form external EMBR. The selectively permeable membrane plays an important role in separation of organic and inorganic pollutants for EMBR. In previous literature, polydimethylsiloxane (PDMS)-related non-porous membranes were mainly used as the selectively permeable membranes, either non-porous PDMS tubular membranes [4,12] or non-porous PDMS flat membranes with porous support [5,6]. However, these membranes usually suffer from poor separation efficiency (non-porous PDMS tubular membranes) or complex fabrication (non-porous PDMS flat membranes with porous support). On the other hand, these membranes, especially non-porous PDMS tubular membranes, are not suitable to use in external EMBR as they are initially designed for internal EMBR facing wastewater and microorganisms together. For this reason, a specially designed membrane for external EMBR with highly permeable to phenol and impermeable to water and salt via a simple fabrication method is needed. Due to controllable thickness, controllable pore size, low mass transfer resistance, high surface hydrophobicity and high mass transfer flux of electrospun membranes, electrospinning gives rise to the interest in EMBR application [13]. In our previous study, an electrospun porous PDMS/PMMA (polymethyl methacrylate) membrane revealed remarkable phenol mass transfer coefficient of 6.7 × 10−7 m/s and salt rejection above 99.9% [14], which indicated its potential in selective separation of phenol and salt. However, electrospun membranes are usually used as the porous support in previous studies rather than selectively permeable membrane directly [5,6]. This was probably attributed to the porous structure of electrospun membranes, which was likely to result in excessive biofilm formation, low membrane separation efficiency and even membrane damage [15,16]. From above discussion on external-installed membrane configuration and selectively permeable membrane for EMBR, it could be concluded that the weaknesses of electrospun membranes in internal EMBR could be overcome when they are used in external EMBR. Consequently, biofouling would be completely avoided without compromising the high-performance phenol mass transfer and salt rejection. To the best of our knowledge, no studies have yet been done on this novel external EMBR system.
2. Materials and methods 2.1. Electrospun membrane PDMS (Sylgard®184, Mw = 60,000 g/mol, Dow Corning), PMMA (182265, Mw = 996,000 g/mol, Sigma-Aldrich), N,N-dimethylformamide (DMF, analytical grade, Sinopharm) and tetrahydrofuran (THF, analytical grade, Sinopharm) were used as received. Electrospinning solution of 10% PDMS and 10% PMMA was prepared in a binary mixture of DMF and THF (weight ratio: 1: 1) by stirring at 50 °C. An electrospinning machine (SS-2535H, Ucalery) was used to fabricate the membrane at a temperature of 25 °C and relative humidity of 50%. The prepared electrospinning solution was loaded into a plastic syringe with a stainless steel spinneret (inner diameter = 0.6 mm). The grounded rotating collector with a 50 rpm speed was placed 16 cm from the spinneret tip. The potential difference between rotating collector and spinneret was 11.5 kV. Morphology, roughness, pore size, composition and wettability of obtained membranes were measured using the methods described in our previous publication [14]. 2.2. External EMBR design The experimental set-up of external EMBR is shown in Fig. 1. The electrospun PDMS/PMMA membrane (effective area: 0.0048 m2) was installed into the membrane cell to divide the external EMBR into feed solution (FS, 500 mL) and receiving solution (RS, 5,000 mL). Firstly, RS was continuously pumped into the membrane cell for phenol permeation. Then, the formed influent (RS and permeated phenol from FS) was continuously pumped into bioreactor (4,500 mL) for biodegradation. During operation, temperature, dissolved oxygen (DO) and pH values in bioreactor were maintained between 24 and 26 °C, 0.5–1.5 mg/L and 7.1–7.5, respectively. HRT of bioreactor influent was adjusted by the peristaltic pump for RS. FS was synthetic phenol-laden saline wastewater, which contained 2.5–20.0 g/L phenol with 10.0 g/L NaCl. RS was in fact a synthetic nutrient solution, which mainly contained 0.2 g/L NH4Cl (Table S1). Acclimated activated sludge with a mixed liquor suspended solids (MLSS) of 4.0 g/L was used as the initial inoculum (named as stage 0), which had already been gradually cultivated under 0.2–0.0 g/L glucose and 0.0–0.2 g/L phenol. The operation period can be divided into 8 stages, from 1 to 8, based on the different phenol concentrations in FS (2.5, 5.0, 7.5, 10.0, 12.5, 15.0, 17.5 and 20.0 g/L). The FS was renewed every day and the membrane was changed for every stage. 2.3. Toxicity analysis According to the standard toxicity measurement of USEPA, the toxicity was evaluated by microtox toxicity test using luminescent bacteria. Luminescent bacteria (Aliivibrio fischeri) were firstly activated at 4 °C using a resuscitation solution. Sample mixtures contained 500 µL of diluted (1000 times) bioreactor influent/effluent, 50 µL of osmotic pressure conditioning solution and 450 µL of activated bacteria solution. Negative quality control solution (distilled water, 500 µL) served 2
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Fig. 1. Schematic of external EMBR system: (1) wastewater, (2) weight balance, (3) peristaltic pump, (4) flowmeter, (5) membrane module, (6) nutrient solution, (7) conductivity meter, (8) bioreactor influent (nutrient solution with phenol), (9) bioreactor, (10) heater and stirrer, (11) aeration, (12) bioreactor effluent.
2.5. Chemical analysis
as control. After 15 min exposure, the toxicity values of water samples to Aliivibrio fischeri were recorded by a Microtox toxicity analyzer (EcloxTM, Hach) as the relative luminescence unit (RLU, rel.unit). Toxicity evaluation of each sample was conducted in triplicate. The luminescence inhibition ratio (LIR, %) was expressed as follows [17]:
LIR =
RLU 0 − RLU × 100% RLU 0
Phenol and ammonium concentrations in FS, bioreactor influent and effluent were measured in triplicate by spectrophotometer (UV-1 800, Mapada) to analyze phenol permeation and removal. Total organic carbon (TOC) concentrations in bioreactor influent and effluent were measured in triplicate by TOC analyzer (Multi N/C 3100, Analytikjena) to evaluate the removal of phenolic pollutants. Conductivities in FS and bioreactor influent were detected via a conductivity meter (DDSJ-308A, Inesa) to analyze salt rejection (SR, %) as follows:
(1)
where RLU0 and RLU represent the average values in control and experimental group, respectively.
SR = (1 − 2.4. Bacterial response
ΔC i ) × 100% Cf
(2)
where ΔC represents the conductivity variation in bioreactor influent, Cf represents the initial conductivity in FS. i
Genomic DNA of sludge samples during different stages (0, 2, 4, 6, 8) was extracted by Soil DNA Kit (e.Z.N.A.™, Omega Bio-Teh) according to the manufacturer’s protocol. The high-throughput sequencing was operated on an Illumina Miseq PE 300 platform for microbial community variation analysis. Four functional genes were applied to evaluate the activities of pollutant removal via qPCR on LightCycler® 480 II (Roche). Specifically, these were (i) phe for phenol hydroxylase; (ii) amoA for ammonium monooxygenase; (iii) narG for nitrate reductase; and (iv) nirS for cd1-containing nitrite reductase. Concrete primers and procedures are described in Table 1. EPS of sludge samples during different stages (0, 2, 4, 6 and 8) was extracted by heat method, and measured by modified Lowry and anthrone-sulfuric method for protein (PN) and polysaccharide (PS) contents. More details were described in our previous study [18].
3. Results and discussion 3.1. Membrane characterization Characterization of the electrospun PDMS/PMMA membrane is described in Fig. S1. The formed membrane existed in a hierarchical non-woven microstructure, with surface roughness of 5.6 ± 0.5 µm. The mean pore size, fiber diameter and thickness were 4.6 ± 0.2, 1.2 ± 0.2 and 76.6 ± 2.3 µm, respectively. The membrane was mainly composed of C (56.2%), O (24.6%) and Si (19.2%) (H is negative in XPS). Measured contact angles for water and phenol on membrane surface were 160.9 ± 2.2° and 0.0°, indicating this membrane was superhydrophobic and superorganophilic. When wastewater (2.5 g/L phenol and 10.0 g/L NaCl) was dropped on membrane surface, a contact angle of 141.5 ± 1.6° was observed. This probably suggested 3
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Table 1 Primers and programs of target genes used in qPCR analysis. Target genes
Primers
Primer sequence (5′-3′)
Programs
phe
pheF pheR amo598F amo718R 1960m2f 2050m2r nirScd3aF nirSR3cd
GTGCTGAC(C/G)AA(C/T)CTG(C/T)TGTTC CGCCAGAACCA(C/T)TT(A/G)TC GAATATGTTCGCCTGATTG CAAAGTACCACCATACGCAG TAYGTSGGGCAGGARAAACTG CGTAGAAGAAGCTGGTGCTGTT GTSAACGTSAAGGARACSGG GASTTCGGRTGSGTCTTGA
Pre-heating (50 °C, 2 min), pre-denaturation (60 °C, 60 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (56 °C, 45 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (58 °C, 45 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (57 °C, 30 s), extension (72 °C, 30 s)
amoA narG nirS
that the phenol in wastewater droplet be adsorbed by this superhydrophobic and superorganophilic membrane instantly when water was rejected. In theory, phenol could be adsorbed by PDMS and PMMA via the formed hydrogen bond [19]. The -Si-O- backbone of PDMS was in turn beneficial for the phenol diffusion [20]. Meanwhile, NaCl was rejected by the superhydrophobic membrane and still remained in FS. Therefore, the electrospun PDMS/PMMA membrane proved to be suitable for phenol separation from saline wastewater.
(95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing
Table 3 Phenol permeation and salt rejection of external EMBR during different stages. Stage
1 2 3 4 5 6 7 8 Control
3.2. Operation of external EMBR 3.2.1. HRT As an integrated system, HRT of EMBR was decided by RS flow rate. Three HRTs (12, 24 and 48 h) were selected, and their effects on phenol permeation and salt rejection for the same FS (5.0 g/L phenol and 10.0 g/L NaCl) are shown in Table 2. With HRT increase, permeated phenol increased from 15.4 ± 3.7 (12 h HRT) to 48.7 ± 4.9 (24 h HRT) and 76.7 ± 6.8 mg/L (48 h HRT). It was clear that a positive correlation was established between HRT and permeated phenol in bioreactor influent. Also, it was interesting to note that no conductivity increase was observed in bioreactor influent under three HRTs, confirming that the membrane remains stable salt rejection performance. This could be due to the superhydrophobic nature of membrane surface which can avoid the invasion of water-soluble salt originating from the FS. In conclusion, longer HRT of wastewater achieved higher phenol permeation. However, too long HRT would result in low phenol permeation rate. For example, the phenol permeation rate of 2.0 mg/L/h under 24 h HRT was higher than that of 1.6 mg/L/h under 48 h HRT. As a result, corresponding lower phenol removal rate and longer wastewater treatment time occurred under 48 h HRT. Therefore the moderate HRT condition of 24 h was used in the subsequent experiment. In EMBR, the selectively permeable membrane achieved phenol permeation and water rejection based on solution-diffusion mechanism under pressure-free condition, which is different from the pore sieving mechanism of nanofiltration, reverse osmosis and forward osmosis in phenol rejection and water permeation under high pressure condition [21–23]. Consequently, the pore blockage could be avoided during EMBR operation. In addition, the big size of membrane pore (4.6 µm), small size of salt ions (< 0.2 nm) and superhydrophobic/superorganophilic membrane surface (160.9°/0° for water/phenol contact angles) reduced the risk of pore blockage. Details of membrane fouling and stability after long-term operation were described in our previous publications [14,15]. Results showed that the morphology, composition
Property of FS
Property of bioreactor influent
Phenol (mg/L)
NaCl (g/L)
Conductivity (ms/cm)
Phenol (mg/L)
Conductivity (ms/cm)
2500 5000 7500 10,000 12,500 15,000 17,500 20,000
10.0 10.0 10.0 10.0 10.0 10.0 10.0 10.0
11.5–11.6 11.5–11.6 11.5–11.6 11.5–11.6 11.6–11.7 11.6–11.7 11.6–11.7 11.6–11.7
14.1 ± 2.7 48.7 ± 4.9 77.3 ± 6.8 88.2 ± 4.6 149.1 ± 9.8 217.5 ± 11.7 243.0 ± 7.9 290.7 ± 10.4 0 (RS)
~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0
and water/phenol contact angles of used electrospun PDMS/PMMA membrane were almost unchanged, suggesting that no pore blocking occurred [14,15]. Moreover, the membrane surface was still superhydrophobic and superorganophilic to guarantee the efficient separation of phenol from water during long-term operation [24–26].
3.2.2. Phenol concentration Table 3 summarizes the detailed compositions of FS and the variations of phenol and conductivities in bioreactor influent. Phenol concentrations in FS were gradually increased from 2.5 g/L in stage 1 to 20.0 g/L in stage 8 with an increment of 2.5 g/L for each stage. In the first stage, 14.1 ± 2.7 mg/L of phenol permeated into RS. Then, phenol concentration in RS gradually increased to 290.7 ± 10.4 mg/L in stage 8 with the phenol level in FS reaching 20.0 g/L. These results indicated that permeated phenol concentrations in RS were positively correlated to initial phenol concentrations in FS. As seen in Fig. 2, a linear relationship between phenol concentrations in RS and FS was established with R2-value of 0.974. In addition, the conductivities in RS were detected at around 2.0 ms/cm for all stages, indicating stable salt rejection performance. The permeation of phenol usually causes conductivity increase in RS [14]. However, this rise in conductivity was negligible when compared with the high conductivity of RS. In conclusion, this membrane demonstrated stable phenol separation from different phenol-laden saline wastewaters, and bioreactor influent mainly contained ammonium (52.0 mg/L) from RS and permeated phenol (14.1–290.7 mg/L) from FS during stages 1–8.
Table 2 Effect of HRT on phenol permeation and salt rejection of external EMBR. HRT (h)
12 24 48
Flow rate
Property of FS
Property of bioreactor influent
FS (L/h)
RS (L/h)
Phenol (g/L)
NaCl (g/L)
Conductivity (ms/cm)
Phenol permeation (mg/L)
Conductivity (ms/cm)
0.188 0.188 0.188
0.375 0.188 0.094
5.0 5.0 5.0
10.0 10.0 10.0
16.5 16.5 16.5
15.4 ± 3.7 48.7 ± 4.9 76.7 ± 6.8
~2.0 ~2.0 ~2.0
4
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had been acclimated for phenol removal. During the following 5 stages (stages 2–6), the effluent phenol concentrations were still maintained at 0 mg/L in the presence of 48.7–217.5 mg/L influent phenol. This implied that phenol proved to be unsatisfactory as both carbon and energy source for denitrifiers and other heterotrophs [27], when a relatively excessive nitrogen source was provided. After day 24, the phenol concentrations in effluent fluctuated between 0.0 and 2.3 mg/L in stages 7 and 8 (influent phenol 243.0–290.7 mg/L), when the phenol removal efficiencies fluctuated between 99.2% and 100.0%. These fluctuations indicated that the phenol of 290.7 mg/L might be the highest concentration, which could be degraded thoroughly and efficiently. If the phenol concentrations in influent further increased, more significant decrease on the phenol removal efficiencies could be expected. To make sure all permeated phenol could be degraded efficiently, no higher phenol dosage in FS was used during the experiment. The removal rate of 290.7 mg/L/d was estimated to be the best that this external EMBR could achieve in terms of removing phenol. In conclusion, the activated sludge in external EMBR exhibited high removal efficiencies for phenol (99.2–100.0%) during stages 1 to 8. Long-term exposure to phenol had no adverse effects on removing pollutants in external EMBR. As the sole carbon and energy source for denitrifiers, the supply of phenol would facilitate the removal of ammonium.
Fig. 2. Relationship between the permeated phenol concentration in RS and initial phenol concentration in FS.
3.3. Pollutant removal in external EMBR
3.3.2. Ammonium As the nitrogen source for microorganisms, ammonium was simultaneously removed in EMBR. Fig. 3 (a) demonstrated that high ammonium removal efficiencies (30.4–38.0%) were achieved immediately in the first 2 days, which implied that the inoculated activated sludge had been acclimated for biological nitrogen removal. However, low removal efficiencies of ammonium (1.0–9.3%) were observed during the subsequent days 3 to 8. This phenomenon indicated that the sludge still needs to adapt to the conditions as the inhibition of phenol on nitrification process occurred even at low concentrations [28]. After adaptation, the ammonium removal efficiencies gradually increased from day 8 to day 32, and finally remained at 83.6% in stage 8, which was remarkably higher than that in initial stage. With reference to the high removal efficiencies of phenol (almost 100.0%), the main limitation regarding nitrogen removal in these stages could be the insufficient supply of carbon source. During the experimental period, nitrite and nitrate were simultaneously produced and removed (Fig. 3(b)), indicating that the nitrogen pathway was nitrification and denitrification. Nitrification usually takes place under aerobic conditions by nitrifiers converting ammonium to nitrate via nitrite. Meanwhile, denitrification is carried out by denitrifiers converting nitrate to dinitrogen gas under anoxic conditions [29]. In this study, the relatively low DO of 0.5–1.5 mg/L facilitated the simultaneous nitrification and denitrification. It is worth noting that the nitrite in effluent significantly increased to 14.7–16.6 mg/L from day 10 to 14. This raised amount of nitrite in effluent without a corresponding nitrate increase indicated that nitrification probably was more sensitive to phenol exposure, which could be attributed to more obvious inhibitory effects of toxicants on nitrifiers than that on denitrifiers and other heterotrophs [30,31]. As a toxicant with negative effect on nitrifiers, phenol increased the possibility of incomplete nitrification and corresponding nitrite accumulation [32,33]. Similarly, Kim et al. also found that nitrification was occasionally upset under phenol exposure, even the sludge was adapted to the wastewater [32].
3.3.1. Phenol As shown in Fig. 3 (a), the average phenol concentrations in effluent were almost 0 mg/L during stage 1, suggesting high phenol removal efficiencies of 100%. This indicated that the inoculated activated sludge
3.3.3. TOC TOC analysis is a simple and effective way to quantify all forms of organic carbon [34]. Besides evaluating the phenol removal performance, TOC values reflect the toxicity of influent and effluent from phenol and its derivatives. Fig. 4 demonstrated that mean influent and effluent TOC concentrations in stage 1 were 23.5 ± 1.0 mg/L and 4.2 ± 1.2 mg/L, respectively. The phenol removal efficiency in stage 1 was 82.2 ± 4.4%. It could be seen that the TOC removal efficiency was
Fig. 3. (a) Phenol and ammonium removal performance of external EMBR during different stages, (b) nitrite and nitrate concentrations in bioreactor influent and effluent during different stages. 5
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LIRs maintained around 22.4 ± 0.2% to 21.9 ± 0.1%, which indicated inferior detoxification performance of 6.3% and 17.0% in stages 1 and 2, respectively. After that, the effluent ammonium clearly decreased and the corresponding effluent LIR reduced from 21.2 ± 0.4% in stage 3 to 16.9 ± 0.5% in stage 8. The resultant detoxification also gradually increased to 70.5% at the end. This suggested that the influent toxicity on activated sludge be remarkably alleviated, which helped to maintain the high pollutant removal efficiencies of EMBR. 3.5. Bacterial response in external EMBR 3.5.1. Microbial community High-throughput sequencing was used to analyze the microbial community response to phenol exposure. Effective sequences of 51,977 to 57,334 were detected in five DNA samples, corresponding with the mean lengths in the range of 438.2–442.5 bp (Table 5). The detected sequences increased from 51,977 in stage 0 to 57,334 in stage 2 when low concentration phenol wastewater was used, and then gradually decreased with the increase of phenol concentration. A further taxonomy analysis of these sequences showed that they could be divided into 117 (stage 0), 212 (stage 2), 222 (stage 4), 195 (stage 6) and 190 (stage 8) OTUs, respectively. All these results suggested that phenol exposure of low concentration could promote the microbial richness and α-diversity, while the high concentration could inhibit the microbial richness and diversity at OTU level. Meanwhile, Table 5 summarizes the variations of Ace, Chao, Shannon and Simpson indices in five samples. These variations also confirmed the phenol effect on bacterial response, as high Ace/Chao represents high richness and high Shannon/ low Simpson represents high α-diversity [38]. The high coverage (> 0.9993) implied that these data reflected the actual microbial community accurately. To further investigate the microbial community variation under phenol exposure, the changed populations at the phylum were identified in Fig. 5 (a). In stage 0, Bacteroidetes occupied the highest percentage of 35.6%. The following major phyla were Actinobacteria (26.7%), Proteobacteria (25.7%) and Firmicutes (9.8%). It seemed that phenol had different impacts on different phyla. The relative abundances of Proteobacteria increased to 31.5–56.1% after phenol exposure. Meanwhile, the relative abundances of Bacteroidetes decreased to 10.3–23.3%. It also should be noted that Saccharibacteria became the dominant phylum in stages 6 and 8 (20.6–35.4%) under high phenol concentration. Phylum Proteobacteria is the main source of nitrifiers and denitrifiers [39,40], and therefore, the high percentages indicated that nitrifiers and denitrifiers might be the main components of microbial community under phenol exposure. As carbon-utilizing bacteria, phylum Saccharibacteria is a common component of activated sludge [41,42]. The increased relative abundance under high phenol concentrations indicated that Saccharibacteria was beneficial for phenol removal in this ecosystem. In stage 8, the Proteobacteria and Saccharibacteria were two dominant phyla with relative abundance of 56.1% and 20.6%, respectively and they might be mainly responsible for the
Fig. 4. TOC removal performance of external EMBR during different stages.
lower than the phenol removal efficiency in the same stage, indicating that complete mineralization of phenol to carbon dioxide had not been achieved. This was consistent with the studies, which also found that TOC still remained in effluent after phenol was removed [35,36]. In the next 7 stages, the influent TOC gradually increased 12.9 fold to 303.1 ± 13.6 mg/L in response to the phenol concentration increase in influent. The consumed TOC gradually increased from 19.3 mg/L in stage 1 to 289.2 mg/L in stage 8. Thus, the effluent TOC only rose slightly to 13.9 ± 0.7 mg/L in stage 8. As a consequence, the TOC removal efficiencies gradually elevated to 95.4 ± 0.2% in stage 8, suggesting the toxicity of influent had almost been alleviated. In the next section, the effect of phenol exposure on toxicity was further investigated using a quantitative analysis. 3.4. Toxicity removal in external EMBR The detoxification performance of EMBR was assessed by the toxicity assessments of individual bioreactor influents and effluents for luminescent bacteria. The relationships of phenol dose – toxicity response in the influent are depicted in Table 4. The average RLU for influent phenol of 14.1 ± 2.7 mg/L was 623.3 ± 1.7 rel.unit. The exposure to higher influent phenol resulted in the decrease in RLUs, which exhibited a RLU of 350.6 ± 1.4 rel.unit under 290.7 ± 10.4 mg/L influent phenol. Based on Table 4, the RLUs of influent were negatively correlated to the concentrations of phenol. In contrast, the LIR rose significantly from 23.9 ± 0.2% to 57.2 ± 0.2% with the phenol increase in influent, indicating that the toxicity of influent significantly increased. The results in 3.3.1 and 3.3.3 showed that high removal efficiencies of phenol and its derivatives were achieved in external EMBR. Therefore, the residual ammonium made the key contribution to LIR as a toxicant for activated sludge [37]. In the first 8 days, ammonium concentrations in effluent were relatively high. At this time, the effluent Table 4 Detoxification performance of external EMBR during different stages. Stage
1 2 3 4 5 6 7 8 Control
Toxicity Influent RLU (rel.unit)
LIR (%)
623.3 603.3 544.4 477.4 422.9 389.7 371.4 350.6 819.4
23.9 26.4 33.6 41.7 48.4 52.4 54.7 57.2 0
± ± ± ± ± ± ± ± ±
1.7 3.4 3.9 9.5 1.0 4.9 1.7 1.4 0.7
± ± ± ± ± ± ± ±
0.2 0.4 0.5 1.4 0.1 0.6 0.2 0.2
6
Effluent RLU (rel.unit)
LIR (%)
635.6 639.6 645.6 643.3 650.9 659.5 672.7 681.1
22.4 21.9 21.2 21.5 20.6 19.5 18.0 16.9
± ± ± ± ± ± ± ±
1.6 0.5 3.3 5.2 4.5 1.7 3.0 4.1
± ± ± ± ± ± ± ±
Detoxification (%) 0.2 0.1 0.4 0.6 0.6 0.4 0.6 0.5
6.3 17.0 36.9 48.4 57.4 62.8 67.1 70.5
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Table 5 Information of DNA samples collected during different stages. Sample
Sequence
Mean length
OTU
Coverage
Ace
Chao
Shannon
Simpson
Seed Stage Stage Stage Stage
51,977 57,334 56,347 56,407 52,178
441.7 441.1 442.5 438.2 441.9
117 212 222 195 190
0.9997 0.9996 0.9996 0.9993 0.9995
125.1 219.2 231.1 215.0 199.3
123.0 218.6 230.1 218.9 197.8
2.5 3.3 3.6 2.9 3.1
0.13 0.08 0.06 0.10 0.08
2 4 6 8
3.5.2. Functional gene To better understand the effect of phenol exposure on pollutant removal and metabolic pathways, the abundance of coding genes for key enzymes in phenol removal and ammonium removal was analyzed. The gene phe encodes phenol hydroxylase as the key enzyme that catalyzes the biotransformation of phenol into catechol for subsequent ring cleavage [43]. The genes amoA, narG and nirS encode ammonium monooxygenase, nitrate reductase and cd1-containing nitrite reductase. These are the catalysts in the oxidation of NH4+ to NO2− in nitrification [44], and the reductions of NO3− to NO2− and NO2− to NO in denitrification, respectively [45]. As shown in Fig. 6, the initial abundance levels of phe, amoA, narG and nirS genes in stage 0 were 2.2E10 ± 5.1E9, 1.1E5 ± 7.7E3, 3.5E7 ± 1.1E6 and 1.9E7 ± 3.6E5 copies/mL DNA, respectively. It should be noted that the inoculated activated sludge had been placed in the reactor before experiment began, without providing nutrient. Therefore, the relevant gene enumerations might be lower than the actual data. The exposure to 48.7 ± 4.9 and 88.2 ± 4.6 mg/L phenol in stages 2 and 4 caused an obvious increase in the genes’ abundance. In stage 4, concentrations of 1.2E11 ± 2.3E9, 8.2E6 ± 1.2E5, 2.3E8 ± 7.6E6 and 1.2E8 ± 4.6E6 copies/mL DNA were detected on phe, amoA, narG and nirS, respectively. No inhibitory effect on functional genes was observed due to the complete removal of phenol. High abundance of functional genes was partly contributed by high enzyme activities, which could accelerate the catalytic behaviors. These findings confirmed that a low concentration of phenol contributed to nitrogen removal in simultaneous nitrification and denitrification as the sole carbon source. This outcome was also underpinned by the observed performance in removing pollutants. Nevertheless, the abundance of phe, amoA, narG and nirS distinctly decreased by 73.9%, 31.2%, 62.6% and 39.1% in stage 6 when the phenol concentration rose from 88.2 ± 4.6 to 217.5 ± 11.7 mg/L. Not only does the nitrification process become inhibited at high phenol concentrations [28], so does the denitrification process [46]. Observations undertaken at the genomic level recorded decreased gene enumeration results. After adaptation, the enumeration numbers gradually increased/recovered to 7.5E10 ± 8.7E9 for phe, 6.7E6 ± 6.1E4 for amoA, 1.5E8 ± 7.4E6 for narG and 1.1E8 ± 5.8E6 copies/mL for nirS in stage 8. Four coding genes were detected in all stages with high abundance levels, confirming that ammonium was mainly removed via nitrification and denitrification using phenol as the sole carbon source. 3.5.3. EPS EPS is the high-molecular-weight secretions of microorganisms, which mainly contains PN and PS. As a complex matrix of microbial flocs, EPS protects the microorganisms from the toxicity of wastewaters [47]. Furthermore, EPS can adsorb organic substances and influence the subsequent mass transfer into microorganisms [48]. Fig. 7 summarizes the quantities of PN and PS extracted from different stages. The EPS in stage 0 mainly contained PN of 5.8 ± 0.1 and PS of 18.2 ± 0.9 mg/g SS. When the phenol concentration increased, PN rapidly became the main component of EPS where the quantity approximately tripled to 16.1 ± 1.7 mg/g SS in stage 8. Meanwhile, the PS significantly decreased to 4.2 ± 1.0 mg/g SS in stage 8. As shown in Fig. 7, PN secretion was inhibited only by high phenol concentration (243.0 ± 7.9 to 290.7 ± 10.4 mg/L), and the inhibitory effect became
Fig. 5. (a) Microbial community compositions, (b) hierarchical clustering tree at OTU level of bacterial samples during different stages.
nitrification-denitrification and carbon consuming process. Fig. 5(b) describes the hierarchical clustering tree at OTU level, which shows the microbial community differences of sludge samples (βdiversity) in a visual way. It was clear that overall taxonomic composition and bacterial community structure altered more significantly under higher phenol concentration exposure (stages 6 and 8) than those under low phenol concentration exposure (stages 2 and 4).
7
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Fig. 6. Enumeration of (a) phe, (b) amoA, (c) narG and (d) nirS by qPCR during different stages.
4. Conclusion This study is the first to reveal that this novel external EMBR is feasible to treat phenol-laden saline wastewaters. Phenol separation from saline wastewater was achieved by the superhydrophobic/superorganophilic membrane with contact angles of 160.9°/0° for water/ phenol. Phenol (14.1–290.7 mg/L) and ammonium (0.5–43.5 mg/L) were simultaneously removed in significant amounts and wastewater toxicity (6.3–70.5%) was reduced. Microbial community variation (mainly Proteobacteria and Saccharibacteria) and functional gene enumeration (phe, amoA, narG and nirS) confirmed that ammonium was removed via nitrification-denitrification when phenol acted as the carbon and energy source for denitrifiers and other heterotrophs. Membrane fouling (pore blockage, biofouling) was avoided due to the solution-diffusion permeation mechanism and externally installed membrane configuration. These findings provide insights into the understanding and application of external EMBR mainly for organic-inorganic composite wastewaters.
Fig. 7. EPS qualities (PN and PS) of sludge samples during different stages.
Declaration of Competing Interest
obvious on PS even at low phenol concentration (14.1 ± 2.7 to 217.5 ± 11.7 mg/L). As seen from the variations of PN and PS quantities, PN may be more important in EPS for toxicity resistance, phenol adsorption and transfer during the phenol-laden wastewater treatments. In this study, EPS level (sum of PN and PS) was found to decrease from 24.0 mg/g SS (stage 0) to 20.3 mg/g SS (stage 8). Meanwhile, removed phenol concentration was found to gradually increase as shown in Fig. 3 (a). These results implied that more and more phenol being adsorbed and transferred to the microorganisms under low EPS level [49]. This phenomenon is well documented in another study, which also found that the permeability of microorganisms was lower at higher level of EPS [50].
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
Acknowledgements The authors gratefully acknowledge the support from Major Science and Technology Program for Water Pollution Control and Treatment (2017ZX07202), National Natural Science Foundation of China (No. 21577089 and 21737002), Science and Technology Program of Xuzhou (KC17136). 8
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Appendix A. Supplementary data
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Novel external extractive membrane bioreactor (EMBR) using electrospun polydimethylsiloxane/polymethyl methacrylate membrane for phenol-laden saline wastewater ⁎
Long-Fei Rena, Huu Hao Ngob, Cuina Buc, Chenghao Gec, Shou-Qing Nic, Jiahui Shaoa, , Yiliang Hea a b c
School of Environmental Science and Engineering, Shanghai Jiao Tong University, No. 800 Dongchuan Road, Shanghai 200240, Shanghai, PR China Centre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW 2007, Australia School of Environmental Science and Engineering, Shandong University, No. 26 South Shanda Road, Jinan 250100, Shandong, PR China
H I GH L IG H T S
EMBR was set up with superhydrophobic/organophilic electrospun membrane. • External EMBR achieved the salt rejection, phenol permeation and biodegradation. • External and ammonium were simultaneously removed in external EMBR for detoxication. • Phenol • Microbial community, gene enumeration and EPS release varied with phenol conc.
A R T I C LE I N FO
A B S T R A C T
Keywords: Extractive membrane bioreactor Electrospun membrane Polydimethylsiloxane Phenol removal Bacterial response
Phenol-laden saline wastewaters can adversely affect water, groundwater, soil, organisms and ecosystems. Given that frequently-used biodegradation process is generally inhibited by salinity, this work aims to solve the problem through a novel configuration of external extractive membrane bioreactor (EMBR) for the objective of simultaneous phenol permeation, salt rejection and biodegradation. Contact angles of 160.9 ± 2.2° (water) and 0.0° (phenol) were observed on the electrospun polydimethylsiloxane/polymethyl methacrylate (PDMS/PMMA) membrane, suggesting this superhydrophobic/superorganophilic membrane was suitable for separating phenol from water-soluble salt. Phenol ranging from 14.1 ± 2.7 to 290.7 ± 10.4 mg/L (stages 1 to 8) was continuously permeated and completely biodegraded in external EMBR under a hydraulic retention time (HRT) of 24 h, which corresponded with detoxification performance improving from 6.3% to 70.5%. After phenol exposure of 8 stages, Proteobacteria and Saccharibacteria became the main phyla for microorganisms. Enumeration of functional genes (phe, amoA, narG, nirS) confirmed that phenol was mainly consumed by denitrifiers and other heterotrophs as the sole carbon and energy source via oxidation and ring cleavage. As bacterial responses, these genes’ proliferation was promoted under low phenol concentrations but inhibited under high phenol concentrations. Meanwhile, results of extracellular polymeric substances revealed that protein was the key substance in toxicity resistance, phenol adsorption and transfer.
1. Introduction Phenol-laden saline wastewaters are widely generated in various industries including tanneries, oil refineries and olive oil mills [1]. The coexistence of phenol and salt in wastewater enhances its toxicity and adversely impacts on ecosystem, such as altering aquatic environments and damaging valuable resources [2]. Moreover, effective treatment of
⁎
phenol-laden wastewater becomes more challenging due to the presence of salinity, which is known as serious inhibitory effects on biodegradation for wastewater [3]. Compared with traditional biodegradation process, extractive membrane bioreactor (EMBR) is a novel combined process of membrane separation and biodegradation for wastewater treatment [4]. A hydrophobic-organophilic membrane was immersed in EMBR to divide
Corresponding author. E-mail address: [email protected] (J. Shao).
https://doi.org/10.1016/j.cej.2019.123179 Received 9 August 2019; Received in revised form 11 October 2019; Accepted 14 October 2019 1385-8947/ © 2019 Elsevier B.V. All rights reserved.
Please cite this article as: Long-Fei Ren, et al., Chemical Engineering Journal, https://doi.org/10.1016/j.cej.2019.123179
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Therefore, a novel external EMBR was firstly set up using an electrospun PDMS/PMMA membrane to treat phenol-laden saline wastewater. Then, the effects of operation conditions (hydraulic retention time (HRT) and concentration) were investigated to achieve efficient phenol separation and biodegradation. The objectives of this study were to: (i) investigate the phenol permeation and salt rejection; (ii) examine the pollutant removal and detoxification; (iii) evaluate the bacterial responses (microbial community, functional gene and extracellular polymeric substances (EPS)) under different phenol-laden saline wastewaters and (iv) elucidate the degradation pathway of phenol. This study should provide better understanding of external EMBR to facilitate phenol-laden saline wastewater treatment.
it into wastewater and bioreactor. Organic pollutants in wastewater could transport through the selectively permeable membrane into bioreactor via solution-diffusion for subsequent biodegradation [4,5]. Meanwhile, inorganic pollutants (e.g., salinity, acid, alkali) are separated by the hydrophobic membrane surface, which can probably affect biodegradation efficiency [4,5]. Therefore, bioreactor could be adjusted to suitable condition for the internal microorganisms to accelerate the removal of permeated organic pollutants and provided nutrients (e.g., nitrogen, phosphorus). These attributes make EMBR suitable for the treatment of toxic recalcitrant organic wastewater where harsh conditions exist (e.g., phenol-laden saline wastewater). EMBR has been widely reported in the treatments of phenol saline wastewaters [4–6]. All these attempts on EMBR used the internal-installed membrane configuration, which facilitated the simultaneous phenol separation and phenol biodegradation. Though the units of wastewater, selectively permeable membrane and microorganisms are excessively integrated in this configuration, it is difficulty to control phenol separation or phenol biodegradation precisely. Consequently, a minor problem in wastewater, selectively permeable membrane or microorganisms may collapse the whole EMBR system. Moreover, the internal-installed membrane configuration is prone to make membranes being subjected to biomass adhesion even biofouling from the microorganisms in bioreactor [7–9]. In comparison, the external-installed membrane configuration could address the above issues effectively: 1) relatively independent units of wastewater, selectively permeable membrane and microorganisms are easier to control, 2) adhesion of biomass or colloidal organic matter to the membrane surface would be completely avoided to free membrane from biofouling [10]. Though the external-installed membrane configuration has been successfully used in MBR fields [10,11], this membrane configuration has not been used in EMBR to form external EMBR. The selectively permeable membrane plays an important role in separation of organic and inorganic pollutants for EMBR. In previous literature, polydimethylsiloxane (PDMS)-related non-porous membranes were mainly used as the selectively permeable membranes, either non-porous PDMS tubular membranes [4,12] or non-porous PDMS flat membranes with porous support [5,6]. However, these membranes usually suffer from poor separation efficiency (non-porous PDMS tubular membranes) or complex fabrication (non-porous PDMS flat membranes with porous support). On the other hand, these membranes, especially non-porous PDMS tubular membranes, are not suitable to use in external EMBR as they are initially designed for internal EMBR facing wastewater and microorganisms together. For this reason, a specially designed membrane for external EMBR with highly permeable to phenol and impermeable to water and salt via a simple fabrication method is needed. Due to controllable thickness, controllable pore size, low mass transfer resistance, high surface hydrophobicity and high mass transfer flux of electrospun membranes, electrospinning gives rise to the interest in EMBR application [13]. In our previous study, an electrospun porous PDMS/PMMA (polymethyl methacrylate) membrane revealed remarkable phenol mass transfer coefficient of 6.7 × 10−7 m/s and salt rejection above 99.9% [14], which indicated its potential in selective separation of phenol and salt. However, electrospun membranes are usually used as the porous support in previous studies rather than selectively permeable membrane directly [5,6]. This was probably attributed to the porous structure of electrospun membranes, which was likely to result in excessive biofilm formation, low membrane separation efficiency and even membrane damage [15,16]. From above discussion on external-installed membrane configuration and selectively permeable membrane for EMBR, it could be concluded that the weaknesses of electrospun membranes in internal EMBR could be overcome when they are used in external EMBR. Consequently, biofouling would be completely avoided without compromising the high-performance phenol mass transfer and salt rejection. To the best of our knowledge, no studies have yet been done on this novel external EMBR system.
2. Materials and methods 2.1. Electrospun membrane PDMS (Sylgard®184, Mw = 60,000 g/mol, Dow Corning), PMMA (182265, Mw = 996,000 g/mol, Sigma-Aldrich), N,N-dimethylformamide (DMF, analytical grade, Sinopharm) and tetrahydrofuran (THF, analytical grade, Sinopharm) were used as received. Electrospinning solution of 10% PDMS and 10% PMMA was prepared in a binary mixture of DMF and THF (weight ratio: 1: 1) by stirring at 50 °C. An electrospinning machine (SS-2535H, Ucalery) was used to fabricate the membrane at a temperature of 25 °C and relative humidity of 50%. The prepared electrospinning solution was loaded into a plastic syringe with a stainless steel spinneret (inner diameter = 0.6 mm). The grounded rotating collector with a 50 rpm speed was placed 16 cm from the spinneret tip. The potential difference between rotating collector and spinneret was 11.5 kV. Morphology, roughness, pore size, composition and wettability of obtained membranes were measured using the methods described in our previous publication [14]. 2.2. External EMBR design The experimental set-up of external EMBR is shown in Fig. 1. The electrospun PDMS/PMMA membrane (effective area: 0.0048 m2) was installed into the membrane cell to divide the external EMBR into feed solution (FS, 500 mL) and receiving solution (RS, 5,000 mL). Firstly, RS was continuously pumped into the membrane cell for phenol permeation. Then, the formed influent (RS and permeated phenol from FS) was continuously pumped into bioreactor (4,500 mL) for biodegradation. During operation, temperature, dissolved oxygen (DO) and pH values in bioreactor were maintained between 24 and 26 °C, 0.5–1.5 mg/L and 7.1–7.5, respectively. HRT of bioreactor influent was adjusted by the peristaltic pump for RS. FS was synthetic phenol-laden saline wastewater, which contained 2.5–20.0 g/L phenol with 10.0 g/L NaCl. RS was in fact a synthetic nutrient solution, which mainly contained 0.2 g/L NH4Cl (Table S1). Acclimated activated sludge with a mixed liquor suspended solids (MLSS) of 4.0 g/L was used as the initial inoculum (named as stage 0), which had already been gradually cultivated under 0.2–0.0 g/L glucose and 0.0–0.2 g/L phenol. The operation period can be divided into 8 stages, from 1 to 8, based on the different phenol concentrations in FS (2.5, 5.0, 7.5, 10.0, 12.5, 15.0, 17.5 and 20.0 g/L). The FS was renewed every day and the membrane was changed for every stage. 2.3. Toxicity analysis According to the standard toxicity measurement of USEPA, the toxicity was evaluated by microtox toxicity test using luminescent bacteria. Luminescent bacteria (Aliivibrio fischeri) were firstly activated at 4 °C using a resuscitation solution. Sample mixtures contained 500 µL of diluted (1000 times) bioreactor influent/effluent, 50 µL of osmotic pressure conditioning solution and 450 µL of activated bacteria solution. Negative quality control solution (distilled water, 500 µL) served 2
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Fig. 1. Schematic of external EMBR system: (1) wastewater, (2) weight balance, (3) peristaltic pump, (4) flowmeter, (5) membrane module, (6) nutrient solution, (7) conductivity meter, (8) bioreactor influent (nutrient solution with phenol), (9) bioreactor, (10) heater and stirrer, (11) aeration, (12) bioreactor effluent.
2.5. Chemical analysis
as control. After 15 min exposure, the toxicity values of water samples to Aliivibrio fischeri were recorded by a Microtox toxicity analyzer (EcloxTM, Hach) as the relative luminescence unit (RLU, rel.unit). Toxicity evaluation of each sample was conducted in triplicate. The luminescence inhibition ratio (LIR, %) was expressed as follows [17]:
LIR =
RLU 0 − RLU × 100% RLU 0
Phenol and ammonium concentrations in FS, bioreactor influent and effluent were measured in triplicate by spectrophotometer (UV-1 800, Mapada) to analyze phenol permeation and removal. Total organic carbon (TOC) concentrations in bioreactor influent and effluent were measured in triplicate by TOC analyzer (Multi N/C 3100, Analytikjena) to evaluate the removal of phenolic pollutants. Conductivities in FS and bioreactor influent were detected via a conductivity meter (DDSJ-308A, Inesa) to analyze salt rejection (SR, %) as follows:
(1)
where RLU0 and RLU represent the average values in control and experimental group, respectively.
SR = (1 − 2.4. Bacterial response
ΔC i ) × 100% Cf
(2)
where ΔC represents the conductivity variation in bioreactor influent, Cf represents the initial conductivity in FS. i
Genomic DNA of sludge samples during different stages (0, 2, 4, 6, 8) was extracted by Soil DNA Kit (e.Z.N.A.™, Omega Bio-Teh) according to the manufacturer’s protocol. The high-throughput sequencing was operated on an Illumina Miseq PE 300 platform for microbial community variation analysis. Four functional genes were applied to evaluate the activities of pollutant removal via qPCR on LightCycler® 480 II (Roche). Specifically, these were (i) phe for phenol hydroxylase; (ii) amoA for ammonium monooxygenase; (iii) narG for nitrate reductase; and (iv) nirS for cd1-containing nitrite reductase. Concrete primers and procedures are described in Table 1. EPS of sludge samples during different stages (0, 2, 4, 6 and 8) was extracted by heat method, and measured by modified Lowry and anthrone-sulfuric method for protein (PN) and polysaccharide (PS) contents. More details were described in our previous study [18].
3. Results and discussion 3.1. Membrane characterization Characterization of the electrospun PDMS/PMMA membrane is described in Fig. S1. The formed membrane existed in a hierarchical non-woven microstructure, with surface roughness of 5.6 ± 0.5 µm. The mean pore size, fiber diameter and thickness were 4.6 ± 0.2, 1.2 ± 0.2 and 76.6 ± 2.3 µm, respectively. The membrane was mainly composed of C (56.2%), O (24.6%) and Si (19.2%) (H is negative in XPS). Measured contact angles for water and phenol on membrane surface were 160.9 ± 2.2° and 0.0°, indicating this membrane was superhydrophobic and superorganophilic. When wastewater (2.5 g/L phenol and 10.0 g/L NaCl) was dropped on membrane surface, a contact angle of 141.5 ± 1.6° was observed. This probably suggested 3
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Table 1 Primers and programs of target genes used in qPCR analysis. Target genes
Primers
Primer sequence (5′-3′)
Programs
phe
pheF pheR amo598F amo718R 1960m2f 2050m2r nirScd3aF nirSR3cd
GTGCTGAC(C/G)AA(C/T)CTG(C/T)TGTTC CGCCAGAACCA(C/T)TT(A/G)TC GAATATGTTCGCCTGATTG CAAAGTACCACCATACGCAG TAYGTSGGGCAGGARAAACTG CGTAGAAGAAGCTGGTGCTGTT GTSAACGTSAAGGARACSGG GASTTCGGRTGSGTCTTGA
Pre-heating (50 °C, 2 min), pre-denaturation (60 °C, 60 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (56 °C, 45 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (58 °C, 45 s), extension (72 °C, 30 s) Pre-heating (50 °C, 2 min), pre-denaturation (57 °C, 30 s), extension (72 °C, 30 s)
amoA narG nirS
that the phenol in wastewater droplet be adsorbed by this superhydrophobic and superorganophilic membrane instantly when water was rejected. In theory, phenol could be adsorbed by PDMS and PMMA via the formed hydrogen bond [19]. The -Si-O- backbone of PDMS was in turn beneficial for the phenol diffusion [20]. Meanwhile, NaCl was rejected by the superhydrophobic membrane and still remained in FS. Therefore, the electrospun PDMS/PMMA membrane proved to be suitable for phenol separation from saline wastewater.
(95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing (95 °C, 10 min), denaturation (95 °C, 15 s), annealing
Table 3 Phenol permeation and salt rejection of external EMBR during different stages. Stage
1 2 3 4 5 6 7 8 Control
3.2. Operation of external EMBR 3.2.1. HRT As an integrated system, HRT of EMBR was decided by RS flow rate. Three HRTs (12, 24 and 48 h) were selected, and their effects on phenol permeation and salt rejection for the same FS (5.0 g/L phenol and 10.0 g/L NaCl) are shown in Table 2. With HRT increase, permeated phenol increased from 15.4 ± 3.7 (12 h HRT) to 48.7 ± 4.9 (24 h HRT) and 76.7 ± 6.8 mg/L (48 h HRT). It was clear that a positive correlation was established between HRT and permeated phenol in bioreactor influent. Also, it was interesting to note that no conductivity increase was observed in bioreactor influent under three HRTs, confirming that the membrane remains stable salt rejection performance. This could be due to the superhydrophobic nature of membrane surface which can avoid the invasion of water-soluble salt originating from the FS. In conclusion, longer HRT of wastewater achieved higher phenol permeation. However, too long HRT would result in low phenol permeation rate. For example, the phenol permeation rate of 2.0 mg/L/h under 24 h HRT was higher than that of 1.6 mg/L/h under 48 h HRT. As a result, corresponding lower phenol removal rate and longer wastewater treatment time occurred under 48 h HRT. Therefore the moderate HRT condition of 24 h was used in the subsequent experiment. In EMBR, the selectively permeable membrane achieved phenol permeation and water rejection based on solution-diffusion mechanism under pressure-free condition, which is different from the pore sieving mechanism of nanofiltration, reverse osmosis and forward osmosis in phenol rejection and water permeation under high pressure condition [21–23]. Consequently, the pore blockage could be avoided during EMBR operation. In addition, the big size of membrane pore (4.6 µm), small size of salt ions (< 0.2 nm) and superhydrophobic/superorganophilic membrane surface (160.9°/0° for water/phenol contact angles) reduced the risk of pore blockage. Details of membrane fouling and stability after long-term operation were described in our previous publications [14,15]. Results showed that the morphology, composition
Property of FS
Property of bioreactor influent
Phenol (mg/L)
NaCl (g/L)
Conductivity (ms/cm)
Phenol (mg/L)
Conductivity (ms/cm)
2500 5000 7500 10,000 12,500 15,000 17,500 20,000
10.0 10.0 10.0 10.0 10.0 10.0 10.0 10.0
11.5–11.6 11.5–11.6 11.5–11.6 11.5–11.6 11.6–11.7 11.6–11.7 11.6–11.7 11.6–11.7
14.1 ± 2.7 48.7 ± 4.9 77.3 ± 6.8 88.2 ± 4.6 149.1 ± 9.8 217.5 ± 11.7 243.0 ± 7.9 290.7 ± 10.4 0 (RS)
~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0 ~2.0
and water/phenol contact angles of used electrospun PDMS/PMMA membrane were almost unchanged, suggesting that no pore blocking occurred [14,15]. Moreover, the membrane surface was still superhydrophobic and superorganophilic to guarantee the efficient separation of phenol from water during long-term operation [24–26].
3.2.2. Phenol concentration Table 3 summarizes the detailed compositions of FS and the variations of phenol and conductivities in bioreactor influent. Phenol concentrations in FS were gradually increased from 2.5 g/L in stage 1 to 20.0 g/L in stage 8 with an increment of 2.5 g/L for each stage. In the first stage, 14.1 ± 2.7 mg/L of phenol permeated into RS. Then, phenol concentration in RS gradually increased to 290.7 ± 10.4 mg/L in stage 8 with the phenol level in FS reaching 20.0 g/L. These results indicated that permeated phenol concentrations in RS were positively correlated to initial phenol concentrations in FS. As seen in Fig. 2, a linear relationship between phenol concentrations in RS and FS was established with R2-value of 0.974. In addition, the conductivities in RS were detected at around 2.0 ms/cm for all stages, indicating stable salt rejection performance. The permeation of phenol usually causes conductivity increase in RS [14]. However, this rise in conductivity was negligible when compared with the high conductivity of RS. In conclusion, this membrane demonstrated stable phenol separation from different phenol-laden saline wastewaters, and bioreactor influent mainly contained ammonium (52.0 mg/L) from RS and permeated phenol (14.1–290.7 mg/L) from FS during stages 1–8.
Table 2 Effect of HRT on phenol permeation and salt rejection of external EMBR. HRT (h)
12 24 48
Flow rate
Property of FS
Property of bioreactor influent
FS (L/h)
RS (L/h)
Phenol (g/L)
NaCl (g/L)
Conductivity (ms/cm)
Phenol permeation (mg/L)
Conductivity (ms/cm)
0.188 0.188 0.188
0.375 0.188 0.094
5.0 5.0 5.0
10.0 10.0 10.0
16.5 16.5 16.5
15.4 ± 3.7 48.7 ± 4.9 76.7 ± 6.8
~2.0 ~2.0 ~2.0
4
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had been acclimated for phenol removal. During the following 5 stages (stages 2–6), the effluent phenol concentrations were still maintained at 0 mg/L in the presence of 48.7–217.5 mg/L influent phenol. This implied that phenol proved to be unsatisfactory as both carbon and energy source for denitrifiers and other heterotrophs [27], when a relatively excessive nitrogen source was provided. After day 24, the phenol concentrations in effluent fluctuated between 0.0 and 2.3 mg/L in stages 7 and 8 (influent phenol 243.0–290.7 mg/L), when the phenol removal efficiencies fluctuated between 99.2% and 100.0%. These fluctuations indicated that the phenol of 290.7 mg/L might be the highest concentration, which could be degraded thoroughly and efficiently. If the phenol concentrations in influent further increased, more significant decrease on the phenol removal efficiencies could be expected. To make sure all permeated phenol could be degraded efficiently, no higher phenol dosage in FS was used during the experiment. The removal rate of 290.7 mg/L/d was estimated to be the best that this external EMBR could achieve in terms of removing phenol. In conclusion, the activated sludge in external EMBR exhibited high removal efficiencies for phenol (99.2–100.0%) during stages 1 to 8. Long-term exposure to phenol had no adverse effects on removing pollutants in external EMBR. As the sole carbon and energy source for denitrifiers, the supply of phenol would facilitate the removal of ammonium.
Fig. 2. Relationship between the permeated phenol concentration in RS and initial phenol concentration in FS.
3.3. Pollutant removal in external EMBR
3.3.2. Ammonium As the nitrogen source for microorganisms, ammonium was simultaneously removed in EMBR. Fig. 3 (a) demonstrated that high ammonium removal efficiencies (30.4–38.0%) were achieved immediately in the first 2 days, which implied that the inoculated activated sludge had been acclimated for biological nitrogen removal. However, low removal efficiencies of ammonium (1.0–9.3%) were observed during the subsequent days 3 to 8. This phenomenon indicated that the sludge still needs to adapt to the conditions as the inhibition of phenol on nitrification process occurred even at low concentrations [28]. After adaptation, the ammonium removal efficiencies gradually increased from day 8 to day 32, and finally remained at 83.6% in stage 8, which was remarkably higher than that in initial stage. With reference to the high removal efficiencies of phenol (almost 100.0%), the main limitation regarding nitrogen removal in these stages could be the insufficient supply of carbon source. During the experimental period, nitrite and nitrate were simultaneously produced and removed (Fig. 3(b)), indicating that the nitrogen pathway was nitrification and denitrification. Nitrification usually takes place under aerobic conditions by nitrifiers converting ammonium to nitrate via nitrite. Meanwhile, denitrification is carried out by denitrifiers converting nitrate to dinitrogen gas under anoxic conditions [29]. In this study, the relatively low DO of 0.5–1.5 mg/L facilitated the simultaneous nitrification and denitrification. It is worth noting that the nitrite in effluent significantly increased to 14.7–16.6 mg/L from day 10 to 14. This raised amount of nitrite in effluent without a corresponding nitrate increase indicated that nitrification probably was more sensitive to phenol exposure, which could be attributed to more obvious inhibitory effects of toxicants on nitrifiers than that on denitrifiers and other heterotrophs [30,31]. As a toxicant with negative effect on nitrifiers, phenol increased the possibility of incomplete nitrification and corresponding nitrite accumulation [32,33]. Similarly, Kim et al. also found that nitrification was occasionally upset under phenol exposure, even the sludge was adapted to the wastewater [32].
3.3.1. Phenol As shown in Fig. 3 (a), the average phenol concentrations in effluent were almost 0 mg/L during stage 1, suggesting high phenol removal efficiencies of 100%. This indicated that the inoculated activated sludge
3.3.3. TOC TOC analysis is a simple and effective way to quantify all forms of organic carbon [34]. Besides evaluating the phenol removal performance, TOC values reflect the toxicity of influent and effluent from phenol and its derivatives. Fig. 4 demonstrated that mean influent and effluent TOC concentrations in stage 1 were 23.5 ± 1.0 mg/L and 4.2 ± 1.2 mg/L, respectively. The phenol removal efficiency in stage 1 was 82.2 ± 4.4%. It could be seen that the TOC removal efficiency was
Fig. 3. (a) Phenol and ammonium removal performance of external EMBR during different stages, (b) nitrite and nitrate concentrations in bioreactor influent and effluent during different stages. 5
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LIRs maintained around 22.4 ± 0.2% to 21.9 ± 0.1%, which indicated inferior detoxification performance of 6.3% and 17.0% in stages 1 and 2, respectively. After that, the effluent ammonium clearly decreased and the corresponding effluent LIR reduced from 21.2 ± 0.4% in stage 3 to 16.9 ± 0.5% in stage 8. The resultant detoxification also gradually increased to 70.5% at the end. This suggested that the influent toxicity on activated sludge be remarkably alleviated, which helped to maintain the high pollutant removal efficiencies of EMBR. 3.5. Bacterial response in external EMBR 3.5.1. Microbial community High-throughput sequencing was used to analyze the microbial community response to phenol exposure. Effective sequences of 51,977 to 57,334 were detected in five DNA samples, corresponding with the mean lengths in the range of 438.2–442.5 bp (Table 5). The detected sequences increased from 51,977 in stage 0 to 57,334 in stage 2 when low concentration phenol wastewater was used, and then gradually decreased with the increase of phenol concentration. A further taxonomy analysis of these sequences showed that they could be divided into 117 (stage 0), 212 (stage 2), 222 (stage 4), 195 (stage 6) and 190 (stage 8) OTUs, respectively. All these results suggested that phenol exposure of low concentration could promote the microbial richness and α-diversity, while the high concentration could inhibit the microbial richness and diversity at OTU level. Meanwhile, Table 5 summarizes the variations of Ace, Chao, Shannon and Simpson indices in five samples. These variations also confirmed the phenol effect on bacterial response, as high Ace/Chao represents high richness and high Shannon/ low Simpson represents high α-diversity [38]. The high coverage (> 0.9993) implied that these data reflected the actual microbial community accurately. To further investigate the microbial community variation under phenol exposure, the changed populations at the phylum were identified in Fig. 5 (a). In stage 0, Bacteroidetes occupied the highest percentage of 35.6%. The following major phyla were Actinobacteria (26.7%), Proteobacteria (25.7%) and Firmicutes (9.8%). It seemed that phenol had different impacts on different phyla. The relative abundances of Proteobacteria increased to 31.5–56.1% after phenol exposure. Meanwhile, the relative abundances of Bacteroidetes decreased to 10.3–23.3%. It also should be noted that Saccharibacteria became the dominant phylum in stages 6 and 8 (20.6–35.4%) under high phenol concentration. Phylum Proteobacteria is the main source of nitrifiers and denitrifiers [39,40], and therefore, the high percentages indicated that nitrifiers and denitrifiers might be the main components of microbial community under phenol exposure. As carbon-utilizing bacteria, phylum Saccharibacteria is a common component of activated sludge [41,42]. The increased relative abundance under high phenol concentrations indicated that Saccharibacteria was beneficial for phenol removal in this ecosystem. In stage 8, the Proteobacteria and Saccharibacteria were two dominant phyla with relative abundance of 56.1% and 20.6%, respectively and they might be mainly responsible for the
Fig. 4. TOC removal performance of external EMBR during different stages.
lower than the phenol removal efficiency in the same stage, indicating that complete mineralization of phenol to carbon dioxide had not been achieved. This was consistent with the studies, which also found that TOC still remained in effluent after phenol was removed [35,36]. In the next 7 stages, the influent TOC gradually increased 12.9 fold to 303.1 ± 13.6 mg/L in response to the phenol concentration increase in influent. The consumed TOC gradually increased from 19.3 mg/L in stage 1 to 289.2 mg/L in stage 8. Thus, the effluent TOC only rose slightly to 13.9 ± 0.7 mg/L in stage 8. As a consequence, the TOC removal efficiencies gradually elevated to 95.4 ± 0.2% in stage 8, suggesting the toxicity of influent had almost been alleviated. In the next section, the effect of phenol exposure on toxicity was further investigated using a quantitative analysis. 3.4. Toxicity removal in external EMBR The detoxification performance of EMBR was assessed by the toxicity assessments of individual bioreactor influents and effluents for luminescent bacteria. The relationships of phenol dose – toxicity response in the influent are depicted in Table 4. The average RLU for influent phenol of 14.1 ± 2.7 mg/L was 623.3 ± 1.7 rel.unit. The exposure to higher influent phenol resulted in the decrease in RLUs, which exhibited a RLU of 350.6 ± 1.4 rel.unit under 290.7 ± 10.4 mg/L influent phenol. Based on Table 4, the RLUs of influent were negatively correlated to the concentrations of phenol. In contrast, the LIR rose significantly from 23.9 ± 0.2% to 57.2 ± 0.2% with the phenol increase in influent, indicating that the toxicity of influent significantly increased. The results in 3.3.1 and 3.3.3 showed that high removal efficiencies of phenol and its derivatives were achieved in external EMBR. Therefore, the residual ammonium made the key contribution to LIR as a toxicant for activated sludge [37]. In the first 8 days, ammonium concentrations in effluent were relatively high. At this time, the effluent Table 4 Detoxification performance of external EMBR during different stages. Stage
1 2 3 4 5 6 7 8 Control
Toxicity Influent RLU (rel.unit)
LIR (%)
623.3 603.3 544.4 477.4 422.9 389.7 371.4 350.6 819.4
23.9 26.4 33.6 41.7 48.4 52.4 54.7 57.2 0
± ± ± ± ± ± ± ± ±
1.7 3.4 3.9 9.5 1.0 4.9 1.7 1.4 0.7
± ± ± ± ± ± ± ±
0.2 0.4 0.5 1.4 0.1 0.6 0.2 0.2
6
Effluent RLU (rel.unit)
LIR (%)
635.6 639.6 645.6 643.3 650.9 659.5 672.7 681.1
22.4 21.9 21.2 21.5 20.6 19.5 18.0 16.9
± ± ± ± ± ± ± ±
1.6 0.5 3.3 5.2 4.5 1.7 3.0 4.1
± ± ± ± ± ± ± ±
Detoxification (%) 0.2 0.1 0.4 0.6 0.6 0.4 0.6 0.5
6.3 17.0 36.9 48.4 57.4 62.8 67.1 70.5
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Table 5 Information of DNA samples collected during different stages. Sample
Sequence
Mean length
OTU
Coverage
Ace
Chao
Shannon
Simpson
Seed Stage Stage Stage Stage
51,977 57,334 56,347 56,407 52,178
441.7 441.1 442.5 438.2 441.9
117 212 222 195 190
0.9997 0.9996 0.9996 0.9993 0.9995
125.1 219.2 231.1 215.0 199.3
123.0 218.6 230.1 218.9 197.8
2.5 3.3 3.6 2.9 3.1
0.13 0.08 0.06 0.10 0.08
2 4 6 8
3.5.2. Functional gene To better understand the effect of phenol exposure on pollutant removal and metabolic pathways, the abundance of coding genes for key enzymes in phenol removal and ammonium removal was analyzed. The gene phe encodes phenol hydroxylase as the key enzyme that catalyzes the biotransformation of phenol into catechol for subsequent ring cleavage [43]. The genes amoA, narG and nirS encode ammonium monooxygenase, nitrate reductase and cd1-containing nitrite reductase. These are the catalysts in the oxidation of NH4+ to NO2− in nitrification [44], and the reductions of NO3− to NO2− and NO2− to NO in denitrification, respectively [45]. As shown in Fig. 6, the initial abundance levels of phe, amoA, narG and nirS genes in stage 0 were 2.2E10 ± 5.1E9, 1.1E5 ± 7.7E3, 3.5E7 ± 1.1E6 and 1.9E7 ± 3.6E5 copies/mL DNA, respectively. It should be noted that the inoculated activated sludge had been placed in the reactor before experiment began, without providing nutrient. Therefore, the relevant gene enumerations might be lower than the actual data. The exposure to 48.7 ± 4.9 and 88.2 ± 4.6 mg/L phenol in stages 2 and 4 caused an obvious increase in the genes’ abundance. In stage 4, concentrations of 1.2E11 ± 2.3E9, 8.2E6 ± 1.2E5, 2.3E8 ± 7.6E6 and 1.2E8 ± 4.6E6 copies/mL DNA were detected on phe, amoA, narG and nirS, respectively. No inhibitory effect on functional genes was observed due to the complete removal of phenol. High abundance of functional genes was partly contributed by high enzyme activities, which could accelerate the catalytic behaviors. These findings confirmed that a low concentration of phenol contributed to nitrogen removal in simultaneous nitrification and denitrification as the sole carbon source. This outcome was also underpinned by the observed performance in removing pollutants. Nevertheless, the abundance of phe, amoA, narG and nirS distinctly decreased by 73.9%, 31.2%, 62.6% and 39.1% in stage 6 when the phenol concentration rose from 88.2 ± 4.6 to 217.5 ± 11.7 mg/L. Not only does the nitrification process become inhibited at high phenol concentrations [28], so does the denitrification process [46]. Observations undertaken at the genomic level recorded decreased gene enumeration results. After adaptation, the enumeration numbers gradually increased/recovered to 7.5E10 ± 8.7E9 for phe, 6.7E6 ± 6.1E4 for amoA, 1.5E8 ± 7.4E6 for narG and 1.1E8 ± 5.8E6 copies/mL for nirS in stage 8. Four coding genes were detected in all stages with high abundance levels, confirming that ammonium was mainly removed via nitrification and denitrification using phenol as the sole carbon source. 3.5.3. EPS EPS is the high-molecular-weight secretions of microorganisms, which mainly contains PN and PS. As a complex matrix of microbial flocs, EPS protects the microorganisms from the toxicity of wastewaters [47]. Furthermore, EPS can adsorb organic substances and influence the subsequent mass transfer into microorganisms [48]. Fig. 7 summarizes the quantities of PN and PS extracted from different stages. The EPS in stage 0 mainly contained PN of 5.8 ± 0.1 and PS of 18.2 ± 0.9 mg/g SS. When the phenol concentration increased, PN rapidly became the main component of EPS where the quantity approximately tripled to 16.1 ± 1.7 mg/g SS in stage 8. Meanwhile, the PS significantly decreased to 4.2 ± 1.0 mg/g SS in stage 8. As shown in Fig. 7, PN secretion was inhibited only by high phenol concentration (243.0 ± 7.9 to 290.7 ± 10.4 mg/L), and the inhibitory effect became
Fig. 5. (a) Microbial community compositions, (b) hierarchical clustering tree at OTU level of bacterial samples during different stages.
nitrification-denitrification and carbon consuming process. Fig. 5(b) describes the hierarchical clustering tree at OTU level, which shows the microbial community differences of sludge samples (βdiversity) in a visual way. It was clear that overall taxonomic composition and bacterial community structure altered more significantly under higher phenol concentration exposure (stages 6 and 8) than those under low phenol concentration exposure (stages 2 and 4).
7
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Fig. 6. Enumeration of (a) phe, (b) amoA, (c) narG and (d) nirS by qPCR during different stages.
4. Conclusion This study is the first to reveal that this novel external EMBR is feasible to treat phenol-laden saline wastewaters. Phenol separation from saline wastewater was achieved by the superhydrophobic/superorganophilic membrane with contact angles of 160.9°/0° for water/ phenol. Phenol (14.1–290.7 mg/L) and ammonium (0.5–43.5 mg/L) were simultaneously removed in significant amounts and wastewater toxicity (6.3–70.5%) was reduced. Microbial community variation (mainly Proteobacteria and Saccharibacteria) and functional gene enumeration (phe, amoA, narG and nirS) confirmed that ammonium was removed via nitrification-denitrification when phenol acted as the carbon and energy source for denitrifiers and other heterotrophs. Membrane fouling (pore blockage, biofouling) was avoided due to the solution-diffusion permeation mechanism and externally installed membrane configuration. These findings provide insights into the understanding and application of external EMBR mainly for organic-inorganic composite wastewaters.
Fig. 7. EPS qualities (PN and PS) of sludge samples during different stages.
Declaration of Competing Interest
obvious on PS even at low phenol concentration (14.1 ± 2.7 to 217.5 ± 11.7 mg/L). As seen from the variations of PN and PS quantities, PN may be more important in EPS for toxicity resistance, phenol adsorption and transfer during the phenol-laden wastewater treatments. In this study, EPS level (sum of PN and PS) was found to decrease from 24.0 mg/g SS (stage 0) to 20.3 mg/g SS (stage 8). Meanwhile, removed phenol concentration was found to gradually increase as shown in Fig. 3 (a). These results implied that more and more phenol being adsorbed and transferred to the microorganisms under low EPS level [49]. This phenomenon is well documented in another study, which also found that the permeability of microorganisms was lower at higher level of EPS [50].
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
Acknowledgements The authors gratefully acknowledge the support from Major Science and Technology Program for Water Pollution Control and Treatment (2017ZX07202), National Natural Science Foundation of China (No. 21577089 and 21737002), Science and Technology Program of Xuzhou (KC17136). 8
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Appendix A. Supplementary data
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