Long term effects of Pb2+ on the membrane fouling in a hydrolytic-anoxic-oxic-membrane bioreactor treating synthetic electroplating wastewater

Chemosphere 232 (2019) 430e438

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Long term effects of Pb2þ on the membrane fouling in a hydrolyticanoxic-oxic-membrane bioreactor treating synthetic electroplating wastewater Qiong Wang a, Qinxue Wen a, Zhiqiang Chen a, b, * a b

State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology (SKLUWRE, HIT), Harbin, 150090, PR China School of Civil Engineering, Lanzhou University of Technology, Lanzhou, 730070, PR China

h i g h l i g h t s  Comprehensive long-term Pb2þ impacts on membrane fouling were investigated.  The rate of cake layer fouling and pore blocking varied at different Pb2þ loadings.  The components and MW variation of SCFs changed the pore blocking rate.  Changes in flocs cohesive force and LB-EPS impacted the cake layer fouling rate.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 22 April 2019 Received in revised form 23 May 2019 Accepted 26 May 2019 Available online 27 May 2019

Long-term effects of Pb2þ on the operating performance and membrane fouling of two hydrolyticanoxic-oxic-membrane bioreactors treating synthetic electroplating wastewater were investigated. The COD, NHþ 4 -N and TN removal efficiencies decreased by 5.5%, 10.4% and 7.9% with long-term exposure of 2 mg L1 Pb2þ, while serious decreases achieved 25.4%, 35.0% and 26.2% with 6 mg L1 Pb2þ exposure, respectively. 2 mg L1 Pb2þ mitigated the cake layer fouling rate by 25.4% but increased the pore blocking rate by 69.1%, which was contributed by the increase of low and moderate molecular weight (MW) components in the soluble and colloidal foulants (SCFs). 6 mg L1 Pb2þ accelerated the cake layer fouling rate by 101.1%, but mitigated the pore blocking rate by 6.4% due to the increase of high MW SCFs (especially polysaccharides). Thermodynamic analyses showed that Pb2þ regulated the concentration and protein/polysaccharide ratio of loosely bound extracellular polymeric substances, thus changing the flocs hydrophobicity and aggregation capacity, leading the cake layer fouling rate variation. © 2019 Elsevier Ltd. All rights reserved.

Handling Editor: Y Yeomin Yoon Keywords: Pb2þ Membrane fouling Molecular weight distribution Flocs aggregation capacity Electroplating wastewater

1. Introduction Plating Pb is the process that electrodepositing Pb on the inert substrates to obtain excellent soldering, ductile and corrosion resistant surfaces. Pb plated circuit boards, connectors, valves, bearings, semiconductors, transistors, wire and strip are extensively used in electronic and storage battery manufacture (Rosenstein, 1990). Pb plating industry obtains a fast development with the industrialization prosperity, meanwhile, large amount of Pb containing wastewater are produced in the manufacturing

* Corresponding author. School of Environment, Harbin Institute of Technology, Harbin, 150090, PR China. E-mail address: [email protected] (Z. Chen). https://doi.org/10.1016/j.chemosphere.2019.05.231 0045-6535/© 2019 Elsevier Ltd. All rights reserved.

process, with Pb2þ concentration in the range of 0.01e9.70 mg L1 (Rahman et al., 2016). As a frequently detected heavy metal in environment, Pb is listed in the national list of hazardous substances of China due to its high toxicity (MEE, 2016). The reported median lethal concentrations (LC50, 48 h) of Pb2þ on Daphnia magna (Jang and Hwang, 2018), Ceriodaphnia dubia and Daphnia carinata (Cooper et al., 2009) were 0.15, 0.21 and 0.44 mg L1. You et al. (2011)'s research revealed that 40 mg L1 Pb2þ inhibited both anaerobic phosphate release and aerobic phosphate uptake of polyphosphate accumulating organisms completely in 6 h. Dutka and Kwan (1984) regarded that Pb had the minimal toxicity to activated sludge microorganisms compared to Hg, Cu and Cd through short term tests. Previous reports of Pb toxicity mostly focused on its acute effects but lacked the investigation of long-

Q. Wang et al. / Chemosphere 232 (2019) 430e438

term impact (Yuan et al., 2015a). However, water chemistry, sludge physicochemical properties and microbial community dynamics all would affect the bioaccessibility, bioavailability and toxicity of Pb under long-term exposure and made its toxicity underappreciated (Chiang et al., 2016). Considering the increased industrialization elevated Pb concentration in wastewater over recent years, the evaluation of long-term Pb toxicity on activated sludge is therefore of great practical significance. A satisfactory pollutants removal was obtained in our previous research (Wen et al., 2018) using a hydrolytic-anoxic-oxic membrane bioreactor (H/A/O-MBR) treating electroplating wastewater. However, the high energy demand and low water productivity resulted from membrane fouling was always a hindrance to the application of MBR. Therefore, improved insight into the fouling mechanism is crucial towards increased membrane application. In past decades, about 30% of the MBR researches focused on membrane fouling (Yang et al., 2006). Lots efforts have been put on considerable variation of membrane fouling in response to heavy metals (Aftab et al., 2017; Feng et al., 2013). Heavy metals regulate the fouling from two aspects: (1) heavy metals form metal precipitates and directly affect the fouling potential of the sludge mixture; and (2) heavy metals affect the microbial metabolism and indirectly affect the characteristics of membrane foulants, which consist of sludge flocs, biopolymer clusters, colloids, extracellular polymeric substances (EPSs), soluble microbial products (SMPs), dissolved organic matters and inorganic substrates (Lin et al., 2014). Aftab et al. (2017) observed an increased sludge fouling potential in an osmotic MBR with increase of Cr2þ and Pb2þ concentration. Long-term exposure to 5 mg L1 Ag nanoparticles (NPs) resulted in floc size reduction and accelerated the membrane fouling in an anaerobic-anoxic-oxic membrane bioreactor (Yuan et al., 2015b). Sludge flocs in an anoxic-oxic-membrane bioreactor (A/O-MBR) tended to release more protein-like EPSs in facing 1 mg L1 Cu2þ, however, protein would increase the hydrophobicity of flocs and lead to deterioration in sludge filterability (Li et al., 2015; Wen et al., 2016). Furthermore, Feng et al. (2013) observed Cu2þ deposited in the cake layer and on the membrane surface in an A/O-MBR, this confirmed that heavy metals contributed to inorganic fouling. It was also reported that Pb2þ at toxic concentration would inhibit the enzyme catalytic activity, damage the conformation of nucleic acids, and disrupt the cell membrane (Yuan et al., 2015a), therefore, Pb2þ would certainly have effect on the sludge characteristic and the fouling potential. However, limited information on the membrane fouling mechanisms of Pb2þ in wastewater treatment process is available. In addition, membrane fouling is the result of interactions between membrane and sludge mixture, all characteristics of the membrane and sludge will exert impact on membrane fouling. These characteristics usually change simultaneously and some even show opposing effects, making the quantification of their individual effect difficult. This study aims to explore the long-term Pb2þ impacts on membrane fouling in a H/A/O-MBR treating synthetic electroplating wastewater. The membrane fouling evolution and fouling resistance under Pb2þ exposure were evaluated to clarify the fouling feature. Then the characteristics of the supernatant including foulants concentration, molecular weight (MW) distribution and pore blocking model were analyzed. Moreover, the effects of Pb2þ on the flocs aggregation and cake layer formation were also evaluated through thermodynamic interaction analysis. 2. Materials and methods 2.1. H/A/O-MBR set up and operation The lab-scale H/A/O-MBR used in the research consists of a

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hydrolytic acidification tank (6.15 L), an anoxic tank (2.45 L) and an oxic tank (7.00 L). A module of polyvinylidene fluoride (PVDF) hollow fiber membrane with an effective surface area of 0.2 m2 and a nominal pore size of 0.2 mm was immersed in the aerobic tank. The critical flux of the membrane was 18e20 L m2 h1 and the instantaneous permeate flux was maintained at 7.5 L m2 h1 by an intermittent suction mode with 8 min of suction followed by 2 min of relaxation. When the trans-membrane pressure (TMP) reached 30 kPa, the membrane module was taken out for physical and chemical cleaning. The period that the TMP increased to 30 kPa was determinated as one cleaning cycle. The total hydraulic retention time was 13 h and sludge retention time of the A/O tank was 60 d. No sludge was discharged from the hydrolytic acidification tank. The nitrification liquid recycle ratio from the oxic tank to the anoxic tank was 200%. An air diffuser was located at the bottom of the aerobic tank with aeration rate of 0.1 m3 h1 to keep the dissolved oxygen concentration at 2e4 mg L1. The reactor was inoculated with residual sludge obtained from a local municipal wastewater treatment plant. Before the experiment, the sludge was cultivated for 2 months to reach a steady state. Synthetic electroplating wastewater which contained 300 mg L1 COD, 30 mg L1 NHþ 4 -N, 1 50 mg L1 NO PO33 -N and 3 mg L 4 -P was used as the influent and its composition could be found in our previous work (Wen et al., 2018). 2 mg L1 Pb2þ was dosed to one H/A/O-MBR (abbreviated as RP) for 75 d, which was designated as operating phase I, and its concentration increased to 6 mg L1 for another 75 d, which was designated as operating phase II. The other H/A/O-MBR without Pb2þ addition was operated as the control (abbreviated as RC).

2.2. Membrane fouling resistance calculation To obtain the membrane fouling resistance, wiping the fouled membrane module with a sponge and rinsing with deionized water to remove the cake layer (physical cleaning), the reduced portion of TMP was caused by the cake layer. Then immersing the membrane module in 0.5% NaClO solution and 0.1% HCl solution for 2 h each (chemical cleaning), the reduced portion of TMP was caused by pore blocking. And the remaining portion of TMP was caused by the intrinsic membrane. Membrane fouling resistances are described by Darcy's law (Zhu et al., 2018):

J¼

TMP TMP  ¼
mRt m Rm þ Rc þ Rp

(1)

where, J is the permeate flux, m s1; Rt, Rm, Rc and Rp are the total membrane resistance, the intrinsic membrane resistance, the cake layer resistance and the pore blocking resistance, m1, respectively; and m is the dynamic viscosity of the permeate at 20  C, Pa s.

2.3. Membrane filtration tests In the membrane filtration test, PVDF flat membrane (Haiyan Co. Ltd., China) with nominal pore size of 0.22 mm and an effective surface area of 28.7 cm2 was employed. New membranes were soaked and rinsed in Milli-Q water to remove the preservative materials prior to their use. The constant pressure filtration system consists of a filtration cell (Amicon 8200, Millipore, USA), a pressure supply equipment and a data recording system. Before the filtration test, Milli-Q water was filtered through the membrane until the system reached a steady state. In the filtration test, the pressure was fixed at 10 kPa, the stirring velocity was 250 rpm and each sample was performed in duplicate.

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2.4. Interfacial interaction assessment

replaced by zeta potential, mV; and ε0 εr is the permittivity of the suspending liquid, C2 m2 N. k is calculated using Eq. (10).

The total surface tension gtot consists of two components: an apolar van der Waals surface tension component gLW , and a polar acid-based surface tension component gAB . gAB comprises nonadditive electron donor g and electron acceptor gþ components. Eqs. (2) and (3) are used to calculate gtot , which can be applied for both solid and liquid:

sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi ffi P e2 ni z2i k¼ ε0 εr kT

pffiffiffiffiffiffiffiffiffiffiffiffi

gAB ¼ 2 gþ g

(2)

gtot ¼ gLW þ gAB

(3)

where the unit of g is mJ m2. The surface tension parameters of the þ  foulants (gLW s , gs and gs ) can be calculated through a set of three extended Young's equations (Eq. (4)) by measuring the contact angles using three different probe liquids (diiodomethane, pure water and formamide) with known surface tension parameters (gtot , gLW , gþ and g ) (Hong et al., 2013). l l l l

ð1 þ cosqÞgtot l ¼2

qffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi

LW þ gLW s gl

qffiffiffiffiffiffiffiffiffiffiffiffi

 gþ s gl þ

qffiffiffiffiffiffiffiffiffiffiffiffi þ g s gl

(4)

where q is the contact angle; subscripts s and l represent the solid and the liquid, respectively; The free energy of interaction between two identical surfaces immersed in water ðDGcoh Þ evaluating the surface hydrophobicity/ hydrophilicity is calculated according to Eq. (5) (Hong et al., 2013).

DGcoh ¼ 2

qffiffiffiffiffiffiffiffiffi

gLW w 

qffiffiffiffiffiffiffiffiffi qffiffiffiffiffiffiffiffiffi

gLW s

 gLW s

qffiffiffiffiffiffiffiffiffi 

gLW w

  qffiffiffiffiffiffi qffiffiffiffiffiffiffi  qffiffiffiffiffiffiffi pffiffiffiffiffiffi pffiffiffiffiffiffi ffi pffiffiffiffiffiffiffi þ   þ 2 g þ 2 gþ gþ w 2 gs  gw w w 2 gs  qffiffiffiffiffiffiffiffiffiffiffiffi þ  4 g s gs

(5)

where subscripts s and w represent the sludge flocs and the water, respectively; the unit of DGcoh is mJ m2. Interfacial interactions in aqueous media are described through the extended Derjaguin-Landau-Verwey-Overbeek (XDLVO) theory. The total interfacial energy U XDLVO between sludge flocs consists of three energy components: Lifshitz-van der Waals interfacial energy U LW , Lewis acid-base interfacial energy U AB and electrostatic double layer energy U EL (Zhang et al., 2014). The cohesive free energies between two spherical surfaces are calculated by Eqs. (6)e(9): AB LW U XDLVO ðhÞ ¼ U EL sws sws ðhÞ þ U sws ðhÞ þ U sws ðhÞ

U LW sws ðhÞ

¼

2pRh20

qffiffiffiffiffiffiffiffiffi

 gLW s

(6)

qffiffiffiffiffiffiffiffiffi2

gLW w

(7)

h

  qffiffiffiffiffiffi qffiffiffiffiffiffiffi  h0  h pffiffiffiffiffiffi pffiffiffiffiffiffi ffi  gþ gþ g U AB s  w s  gw exp sws ðhÞ ¼  4pRl

l

(8) 2 U EL sws ðhÞ ¼ 2pε0 εr Rj lnð1 þ expð  khÞÞ

(9)

where the unit of U is KT; subscript sws represent two sludge particles contact with each other in the water; R is the radius of the sludge particles, nm; l is the decay length of AB energy, 0.6 nm; h0 is the minimum separation distance, 0.158 nm; k is the inverse Debye screening length, m1; j is the stern potential and could be

(10)

where e is the electron charge, C; ni is the number concentration of ion i in the bulk solution, m3; zi is the valence of ion i; k is Boltzmann's constant, J K1; and T is absolute temperature, K. 2.5. Analytical methods The supernatant was obtained via centrifuging the sludge mixed liquor at 4000 g for 5 min, then the sludge pellet was removed. The extraction of EPSs was carried out according to the two-step heat extraction method descripted by Li, 2016. The sludge pellet was resuspended to the initial volume with NaCl solution (0.05%, 70  C), and followed by vortex mixing for 1 min. The sludge suspension was centrifuged at 4000 g for 10 min, then the supernatant was filtered through 0.45 mm membrane filter to obtain the loosely bound EPS (LB-EPS). The residual sludge pellet was resuspended to the initial volume with 0.05% NaCl solution, and followed by heating in a 60  C water-bath for 30 min. The cooled sludge suspension was centrifuged at 4000 g for 15 min and the filtered supernatant was tightly bound EPS (TB-EPS). The concentrations of polysaccharide and protein were measured by the phenol-sulfuric acid method and the Lowry method, respectively. The influents and membrane effluents were collected every three days for the 2þ measurement of COD, NHþ concentrations. The 4 -N, TN and Pb sludge mixed liquor in the oxic tank was collected every five days for the measurement of mixed liquor suspended solid (MLSS) and mixed liquor volatile suspended solid (MLVSS). These conventional parameters were analyzed according to the Standard Methods (APHA, 2005). The soluble Pb2þ concentration was detected through an inductively coupled plasma optical emission spectrometer (ICP-OES, Optima 7000DV, Perkin Elmer). The contact angle was measured through a contact angle meter (SL150, KINO Industry, USA), each contact angle value was based on arithmetic mean of fifteen independent measurements. Zeta potential was analyzed through an electrophoretic light scattering spectrophotometer (Nano-Z, Malvern, UK). Molecular weight distribution was determined through a high performance size exclusion chromatograph (HP-SEC, LC-10A, Shimadzu, Japan) with a UV detector (254 nm). A gel column (Ultrahydrogel linear 7.8  300 mm, Waters, USA) with a column temperature of 30  C was used. Phosphate buffer (0.73 g L1 KH2PO4, 1.39 g L1 Na2HPO4$12H2O and 7.1 g L1 Na2SO4) was used as the mobile phase with a flow rate of 0.6 mL min1. The sample injection volume was 50 mL. MW calibration curve was done by using polystyrene sulfonate standards (PSS, American polymer standards) with MW of 1.7, 7.5, 15.5, 41, 140, 500 kDa, respectively, and acetone (MW of 58). 3. Results and discussion 3.1. Pb2þ impacts on the treatment performance The pollutants removal efficiencies were continuously monitored to investigate the effects of Pb2þ on the treatment performance (Table 1). The average COD, NHþ 4 -N and TN removal efficiencies were 81.1 ± 3.8%, 85.3 ± 4.7% and 57.7 ± 3.8% in the whole operation phase, respectively, in RC. The average MLSS and MLVSS were 4.44 ± 0.42 and 3.03 ± 0.17 g L1, respectively. The average effluent COD, NHþ 4 -N and TN concentration were 57.72 ± 12.56, 4.32 ± 1.72 and 32.89 ± 2.84 mg L1, respectively.

Q. Wang et al. / Chemosphere 232 (2019) 430e438 Table 1 System performance of the H/A/O-MBRs. Parameter

Time (d) COD removal (%) NHþ 4 -N removal (%) TN removal (%) 2þ Pb removal (%) MLSS (g L1) MLVSS (g L1)

RC

1e150 81.1 ± 3.8 85.3 ± 4.7 57.7 ± 3.8 d 4.44 ± 0.42 3.03 ± 0.17

RP Phase I

Phase II

1e75 75.6 ± 6.4 74.9 ± 7.4 49.8 ± 5.3 85.5 ± 5.1 4.38 ± 0.50 2.81 ± 0.28

76e150 55.7 ± 5.1 50.3 ± 6.7 31.5 ± 6.0 82.0 ± 6.9 2.68 ± 0.57 1.82 ± 0.35

Both the effluent COD and NHþ 4 -N met the discharge standard for electroplating wastewater in China (80 mg COD L1 and 15 mg 1 NHþ 4 -N L , GB 21900-2008), indicating the high performance of the system. As the influent TN increased to about 80 mg L1, the effluent TN increased to 32.89 ± 2.84 mg L1, which exceeded the emission standard (20 mg TN L1). The pollutants removal efficiencies of RP gradually decreased with the long-term exposure to 2 and 6 mg L1 Pb2þ. The average COD, NHþ 4 -N and TN removal efficiencies decreased by 5.5%, 10.4% and 7.9% in phase I compared to those of RC, respectively. More serious decreases in phase II were achieved 25.4%, 35.0% and 26.2%, respectively. Furthermore, the MLSS and MLVSS of RP decreased by 1.4% and 7.3% in phase I and further decreased by 39.6% and 39.9% in phase II compared to those of RC, respectively. Therefore, the system showed resistance to long-term 2 mg L1 Pb2þ exposure, little effects in COD, NHþ 4 -N and TN removal were observed. However, the long-term exposure to 6 mg L1 Pb2þ resulted deterioration of the effluent in terms of 2þ COD, NHþ resistance 4 -N and TN. Therefore, the H/A/O-MBR had Pb in electroplating wastewater treatment, however, the heavy metals concentration in the influent needs to be carefully controlled. 3.2. Pb2þ impacts on the fouling performance 3.2.1. The development of TMP As TMP is an important indicator of the membrane fouling evolution, its variation was monitored continuously throughout the experiment (Fig. 1). The membrane of RC performed 8 cleaning cycles in the whole operation period, its average TMP rising rate rose slowly from 1.26 kPa d1 (cleaning cycle 1) to 1.84 kPa d1 (cleaning cycle 8). The membrane of RP performed 4 and 6 cleaning cycles in phase I and II, respectively. When RP was continuously

Fig. 1. TMP variation of RC and RP.

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exposed to 2 mg L1 Pb2þ (phase I), the average TMP rising rate increased from 1.49 kPa d1 (cleaning cycle 1) to 2.19 kPa d1 (cleaning cycle 3). The membrane fouling was accelerated with 2 mg L1 Pb2þ addition as its average TMP rising rate was 18.6% (cleaning cycle 1) to 45.6% (cleaning cycle 3) higher than that of RC. Then the average TMP rising rate of RP in cleaning cycle 4 decreased to 1.07 kPa d1 and it was 29.6% lower than that of RC (1.52 kPa d1). This indicated the membrane fouling of RP in cleaning cycle 4 was mitigated. When Pb2þ concentration increased to 6 mg L1 (phase II), the average TMP rising rate rose quickly from 1.65 kPa d1 (cleaning cycle 5) to 2.65 kPa d1 (cleaning cycle 10). The membrane fouling deteriorated continuously as its average TMP rising rate was 11.8% (cleaning cycle 5) to 44.0% (cleaning cycle 10) higher than those of RC. The result indicated that long term exposure to Pb2þ would commonly aggravate the membrane fouling, except for a short term mitigation occurred by the end of 2 mg L1 Pb2þ exposure. The emerging membrane fouling mitigation with long-term 2 mg L1 Pb2þ exposure will be discussed in section 3.3 and 3.4. 3.2.2. Membrane fouling resistance distribution The main causes of membrane fouling are the growth of microorganisms on the membrane pore wall, the formation of gel layer and cake layer on the membrane surface. Therefore, the total membrane fouling resistance consists of cake layer resistance, pore blocking resistance and intrinsic membrane resistance. In order to explore the effects of Pb2þ on the fouling process, the average resistance rising rate (dR/dt) representing the fouling rate is evaluated (as shown in Fig. S1 in the Supplementary Material). In all cleaning cycles, cake layer resistance accounted for more than 83% of the total resistance, while pore blocking resistance contributed less than 7% of the total resistance. Therefore, cake layer fouling was always the main pollution form in the treatment system. For RC, its average total resistance rising rate (Rt-ave) and average cake layer resistance rising rate (Rc-ave) increased from 71.27  1010 and 62.37  1010 m1 d1 to 115.36  1010 and 101.00  1010 m1 d1 with the cleaning cycles decreased from 22 d to 14 d, respectively, while the average pore blocking resistance rising rate (Rp-ave) decreased from 3.79  1010 to 1.71  1010 m1 d1. Similar variation also occurred in RP, its Rt-ave and Rc-ave increased from 82.68  1010 and 71.91  1010 m1 d1 to 113.35  1010 and 99.64  1010 m1 d1, Rp-ave decreased from 3.54  1010 to 2.03  1010 m1 d1 in the first three cleaning cycles of phase I, respectively. In phase II, the Rt-ave and Rc-ave increased from 106.35  1010 and 90.95  1010 m1 d1 to 247.18  1010 and 218.51  1010 m1 d1 with the cleaning cycles decreased from 16 d to 9 d, respectively, while, its Rp-ave decreased from 2.97  1010 to 1.45  1010 m1 d1. It was obvious that the Rtave and Rc-ave increased with the increase of membrane fouling, while, Rp-ave decreased. The cake layer on the membrane surface has the function of pre-filtration, intercepting part of organic matters, thus weakening the blockage of membrane pores. Typical cleaning cycles of RP included cleaning cycle 4 (51e77 d, membrane fouling mitigation) in phase I and cleaning cycle 9 (132e141 d, membrane fouling deterioration) in phase II were selected for the analysis of membrane fouling mechanism under Pb2þ exposure. Corresponding to the operating time, cleaning cycle 4 (61e77 d) and 8 (126e139 d) of RC were also selected as the control. More detailed resistance distributions of typical cleaning cycles were compared and listed in Table 2. In cleaning cycle 4, the cake layer resistance and pore blocking resistance of RC were 12.46  1012 and 0.45  1012 m1, respectively. More serious cake layer fouling and pore blocking were observed in RP as its corresponding resistances were 2.31  1012 and 0.76  1012 m1 higher than those of RC, respectively. However, due to the operating time difference of RP (27 d) and RC (17 d) in cleaning cycle 4. The Rc-ave

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Table 2 The resistances distribution of the membrane foulants. Parameter

RC-cycle 4

RP-cycle 4

RC-cycle 8

RP-cycle 9

Time (d) Rt (1012 m1) Rc (1012 m1) Rp (1012 m1) Rm (1012 m1) Rt-aveb (1010 m1 d1) Rc-ave (1010 m1 d1) Rp-ave (1010 m1 d1)

17 14.78 (100%)a 12.46 (84.3%) 0.45 (3.0%) 1.87 (12.7%) 86.94 73.29 2.65

27 17.73 (100%) 14.77 (83.3%) 1.21 (6.8%) 1.75 (9.9%) 65.67 54.70 4.48

14 16.15 (100%) 14.14 (87.5%) 0.24 (1.5%) 1.77 (11.0%) 115.36 101.00 1.71

10 22.71 (100%) 20.31 (89.4%) 0.16 (0.7%) 2.24 (9.9%) 227.10 203.10 1.60

a b

The data in parentheses represent the percentage of the sub-resistance occupying the total resistance. Rt-ave, Rc-ave and Rp-ave represent the average resistance rising rate of the total membrane fouling, cake layer and membrane pore, respectively.

and Rp-ave of RP were 18.59  1010 m1 d1 lower and 1.83  1010 m1 d1 higher than those of RC (73.29  1010 m1 d1 for the Rc-ave, 2.65  1010 m1 d1 for the Rp-ave), respectively. This indicated that after long-term exposure to 2 mg L1 Pb2þ, the cake layer fouling rate decreased but the pore blocking rate increased. In cleaning cycle 8 of RC, the cake layer resistance and pore blocking resistance were 14.14  1012 and 0.24  1012 m1, respectively, while, RP showed more serious cake layer fouling but lighter pore blocking as its corresponding resistances were 6.17  1012 m1 higher and 0.08  1012 m1 lower than those of RC, respectively. As for the fouling rate, the Rc-ave and Rp-ave of RP were 102.10  1010 m1 d1 higher and 0.11  1010 m1 d1 lower than those of RC (101.00  1010 m1 d1 for the Rc-ave, 10 1 1 1.71  10 m d for the Rp-ave), respectively. Therefore, long term exposure to 6 mg L1 Pb2þ would accelerate the cake layer fouling rate but mitigate the pore blocking rate. As Pb2þ showed different impacts on the pore blocking and cake layer fouling processes, the fouling mechanism were also investigated further in section 3.3 and 3.4. 3.3. Mechanism of pore blocking under Pb2þ exposure 3.3.1. The concentration of SCFs It has been well known that the pore blocking is mostly induced by the soluble and colloidal foulants (SCFs) in the supernatant due to their sticky properties (Ji et al., 2008). Surface properties, molecular size and amount of dominant effective foulants all significantly affect this process (Lin et al., 2014). Usually, polysaccharide and protein in sludge suspension are assumed to be the dominant components that contribute to the supernatant fouling characteristics. Their concentrations were analyzed to evaluate the effects of Pb2þ on the membrane pore blocking (Fig. 2). In cleaning cycle 4, 10.15 ± 0.59 mg L1 polysaccharide and 4.59 ± 0.50 mg L1 protein

Fig. 2. Protein and polysaccharide concentrations of SCFs.

were detected in the SCFs of RC, which were 2.82 and 1.78 mg L1 lower than those of RP, respectively. In cleaning cycle 8, the polysaccharide and protein concentrations of RC were 11.20 ± 0.80 and 6.53 ± 0.16 mg L1, which were 5.95 and 2.79 mg L1 lower than those of RP in cleaning cycle 9, respectively. It seemed that Pb2þ stimulated microbes to release more polysaccharide and protein in the supernatant (especially polysaccharide) and this phenomenon was also observed in the previous CeO2 NPs (Ma et al., 2013) and ZnO NPs (Zhang et al., 2017) toxicity tests. Protein and polysaccharide would increase the hydrodynamic diameter of the nanoparticles and metal ions and promote their aggregation, thus the toxicity resistance of the microbes were improved (Hou et al., 2015). More released polysaccharides and proteins resulted in the pore blocking deterioration after long-term 2 mg L1 Pb2þ exposure. However, the pore blocking did not get worse although protein and polysaccharide contents increased also in the SCFs in long term 6 mg L1 Pb2þ exposure. Chen et al. (2017) had reported similar phenomenon and they considered that the rapidly formed cake layer prevented small particles from entering membrane pores and mitigated the pore blocking. It could be speculated that the size difference of the foulants induced different membrane fouling processes, and this was verified hereinafter. 3.3.2. Molecular weight distribution of SCFs MW distribution of SCFs were detected using size exclusion chromatograph (SEC) (Fig. 3). Based on multi-peak Gaussian fitting, the peak area was integrated with software Origin 9.1 to quantify different MW fractions and their percentages (Fig. 4). Three peaks representing different MW fractions were identified. Peak 1

Fig. 3. MW distribution of SCFs in cycle 4 (a), SCFs in cycle 8/9 (b), permeate in cycle 4 (c) and permeate in cycle 8/9 (d).

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(section 3.3.1 had confirmed that polysaccharide was the dominant component), and this resulted severe pore blocking. While longterm 6 mg L1 Pb2þ exposure induced the increase of high MW SCFs, which resulted in a fast formation of gel layer on the membrane surface before the saturation of membrane pores. Therefore, cake resistance contributed more to the membrane fouling than the pore resistance.

Fig. 4. Percentage of each MW fraction in the SCFs (a) and permeate (b).

(0.3e1 kDa), peak 2 (1e10 kDa) and peak 3 (100e1000 kDa) indicated low, moderate and high MW fractions, respectively. According to the analysis on percentages of different MW fractions in RC and RP, the low, moderate and high MW fractions of RC accounted for 65.8%, 32.9% and 1.3% of the total fractions in cleaning cycle 4, while the corresponding fraction percentages of RP were 57.7%, 38.3% and 4.0%, respectively (Fig. 4a). In cleaning cycle 8, these MW fractions of RC were 61.3%, 34.2% and 4.4%, whereas in cleaning cycle 9, these values of RP were 58.3%, 34.6% and 7.1%, respectively (Fig. 4b). Low and moderate MW fractions of SCFs were always in dominant (low MW fraction > 57%, moderate MW fraction > 32%), while the high MW fraction only took up less than 8% of the total SCFs and Pb2þ induced the increase in moderate and high MW fractions. The increase of moderate MW fraction might because SMPs with MW of 1e10 kDa have the strongest complexation capacity among all SMP fractions (Holakoo et al., 2006), therefore, microbes tended to release more moderate MW fraction to decrease the Pb2þ toxicity. The increase of high MW fraction might because of the increase of cell decay, which was caused by the high Pb2þ concentration and generated more biomass associated products (BAPs) having molecules higher than 10 kDa (Jiang et al., 2010). The analysis of the increment of each MW fraction after Pb2þ addition showed that the peak intensities of low, moderate and high MW fractions increased by 2.8%, 0.8% and 0.1% compared to the total SCFs of RC, respectively (Fig. 3a), under long-term of 2 mg L1 Pb2þ exposure. While the peak intensities of low, moderate and high MW fractions under 6 mg L1 Pb2þ exposure increased by 19.4%, 1.0% and 0.3%, respectively (Fig. 3b). Therefore, Pb2þ induced more SCFs release in the treatment systems. In order to verify the role of each MW fraction playing in pore blocking process, MW distributions of the permeate were analyzed and shown in Fig. 3c and d. The results suggested that the high MW fraction could not pass through the membrane pores but might cover the membrane surface to form a gel layer (special cake layer) (Li et al., 2015), especially the high molecular polysaccharides will stimulate the gel layer formation due to their gelation property (Jang et al., 2013). Less than 13% of low MW fraction and 5% of moderate MW fraction in the SCFs were intercepted by the membrane. Therefore, the low and moderate MW fractions induced pore blocking. Jiang et al. (2010) reported that the supernatant containing high percentage of micro-molecules (MW < 20 kDa) exhibited lower retention percentage but higher blocking resistance. Therefore, the membrane fouling process under Pb2þ exposure can be speculated as the following. Long-term 2 mg L1 Pb2þ exposure induced the increase of low and moderate MW SCFs

3.3.3. The pore blocking model In order to further verify the pore blocking mechanism, membrane filtration tests of SCFs of typical cleaning cycles were proceeded. The experimental data were fit to the Hermia model (as shown in Fig. S2 in the Supplementary Material) with R2 value (Table 3) indicating the model fitness. The Hermia model includes four classic filtration models, that are complete blocking, standard blocking, intermediate blocking, and cake filtration. For RC-cycle 4, the primary fouling mechanism was intermediate blocking (R2 > 0.9922), followed by cake filtration (R2 > 0.9675). This implied that membrane pores were mainly covered by macro-molecules, accumulated foulants formed the gel layer. The flux decline by SCFs of RP-cycle 4 fitted the standard blocking (R2 > 0.9864) best and next the intermediate blocking (R2 > 0.9851). As SCFs are heterogeneous in MW distribution and hydrophilicity, the hydrophilic micro-molecules entered membrane pores and induced volume constriction (Liu et al., 2017), while macro-molecules covered membrane pores. For RC-cycle 8, both cake filtration (R2 > 0.8878) and intermediate blocking (R2 > 0.8242) contributed to pore blocking. While for RP-cycle 9, only cake filtration (R2 > 0.8486) fit well the flux decline, this confirmed the fast-forming gel layer. Therefore, cake filtration is the primary pore blocking mechanism in the severe membrane fouling cycle. Such pore blocking models also verified the speculation in section 3.3.2. 3.4. Mechanisms of floc adhesion and cake layer formation under Pb2þ exposure 3.4.1. Surface properties of the flocs Floc adhesion and cake layer formation can be considered as the second stage of membrane fouling in MBRs (Lin et al., 2014), solutes, colloids, EPSs, flocs and inorganic particles jointly construct the cake layer. Flocs are considered a major contributor to cake layer resistance once the cake is formed on the membrane surface. In the cake layer formation stage, sludge flocs get close to the membrane under the hydrodynamic drag forces and adhere on the surface under the thermodynamic forces (Zhang et al., 2014). As the thermodynamic interactions are regarded as the major forces controlling sludge adhesion process, the thermodynamic analysis would provide insight into the fouling behavior under the exposure of Pb2þ. Sludge flocs samples taken from the cleaning cycle 4 and cleaning cycle 8/9 were used to evaluate their thermodynamic interactions. Detected contact angles and zeta potentials, together with calculated interfacial energies are listed in Table 4. The cohesive free energy quantitatively characterizes the degree of hydrophobicity and hydrophilicity, a negative DGcoh means the material surface has hydrophobic property and higher absolute DGcoh value indicates stronger hydrophobicity (Li, 2016). In cleaning cycle 4, the DGcoh of RC and RP were 48.68 and 42.30 mJ m2, respectively. The absolute DGcoh value of RP was 13.1% lower than that of RC, implying the weaker hydrophobicity of sludge flocs in RP after long-term 2 mg L1 Pb2þ exposure. In cycle 8/9, the DGcoh of RC and RP were 51.07 and 59.01 mJ m2, respectively. The absolute DGcoh value of RP was 15.5% higher than that of RC, indicating a stronger hydrophobicity of sludge flocs in RP after long-term 6 mg L1 Pb2þ exposure. Sludge flocs with strong hydrophobicity were more easily adhering

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Table 3 R2 values of regression analyses of pore blocking mechanism. Parameter

Complete blocking

Standard blocking

Intermediate blocking

Cake filtration

RC-cycle 4 RP-cycle 4 RC-cycle 8 RP-cycle 9

0.8987 0.9669 0.7068 0.5957

0.9630 0.9864 0.7730 0.6728

0.9922 0.9851 0.8242 0.7418

0.9675 0.9457 0.8878 0.8486

Table 4 Surface thermodynamic parameters of the sludge flocs. Parameter

Contact angle (o)

RC-cycle 4 RP-cycle 4 RC-cycle 8 RP-cycle 9

Pure water 73.31 ± 1.55 70.48 ± 1.71 74.56 ± 2.12 78.09 ± 2.43

Formamide 45.70 ± 3.62 44.83 ± 2.65 47.01 ± 2.04 53.16 ± 3.16

Diiodomethane 37.66 ± 2.75 38.49 ± 3.38 37.35 ± 1.50 36.11 ± 2.32

Zeta potential

Particle size

DGcoh

（mV） 12.90 ± 0.52 14.44 ± 0.37 11.20 ± 0.65 10.47 ± 0.81

(mm) 58.13 ± 2.61 67.32 ± 1.22 53.07 ± 3.07 40.52 ± 0.70

(mJ m2) 48.68 42.30 51.07 59.01

on the membrane surface and causing cake layer formation (Tian et al., 2015). The effects of Pb2þ on flocs hydrophobicity resulted in the variation of cake layer fouling rate. Such results also confirmed the speculation in section 3.3.2 that the cake layer formation was accelerated under 6 mg L1 Pb2þ exposure.

flocs size in different cleaning cycles (as shown in Table 4) were in direct proportion to the total interfacial interaction energy. Small size flocs would lead serious cake layer fouling (Ji et al., 2008; Lin et al., 2014), which means flocs with lower total interfacial interaction energy would lead more serious cake layer fouling.

3.4.2. Aggregation properties of the flocs According to the XDLVO theory, the interfacial interactions between the sludge flocs can be regarded as the interactions between the spheres. Fig. 5 depicts the total interaction energy variation with separation distance of sludge flocs. Lifshitz-van der Waals (LW), Lewis acid-base (AB) and electrostatic double layer (EL) interfacial energy together constituted the total interfacial energy (Fig. 5a). The LW and AB interfacial interactions are always attractive while the EL interfacial interaction is consistently repulsive, they all increase with the decrease of distance. Debye length represents the spatial scale of the EL interaction. Due to the high ion concentration of electroplating wastewater and high ion strength (0.20 mol L1) of the sludge mixture in electroplating wastewater treatment system, a short Debye length (0.69 nm) was resulted in, which indicated a short-ranged EL interaction between flocs. And this generated an ascendant attractive tendency of the total interaction energy with the decrease of the distance. Fig. 5b showed that the total interaction energy required for cells to be desorbed from sludge flocs were in the order of RP-cycle 4 > RC-cycle 4 > RC-cycle 8 > RP-cycle 9. The higher external energy is required to disperse the floc structure, the stronger aggregation capacity of the sludge flocs would have (Li, 2016). The aggregation capacity of the sludge flocs directly impacts the size of flocs. It is obvious that the average

3.4.3. The components of LB-EPS EPSs, biopolymers originated from microorganisms, can be further classified into LB-EPS and TB-EPS. LB-EPSs locate at the out layer of the flocs and directly contact with the membrane surface during the filtration process. Sludge flocs are three-dimensional matrix in which microbes embed in EPSs, their surface characteristics depend on the properties of LB-EPSs to a large extent (Lin et al., 2014). In order to clarify the functional mechanism of Pb2þ on regulating the flocs characteristics, the dominant constituents including proteins and polysaccharides of LB-EPSs were analyzed (Fig. 6). In cleaning cycle 4 of RP, the average protein concentration of LB-EPS was 9.19 ± 1.68 mg g1VSS and it was 1.29 mg g1VSS lower than that of RC (10.48 ± 0.65 mg g1VSS), while their average polysaccharide concentrations had a negligible difference (3.05 ± 0.43 and 3.02 ± 0.33 mg g1VSS of RP and RC, respectively). The ratio of protein/polysaccharide (PN/PS) of RP was 3.46, and it was 0.45 lower than that of RC. In cleaning cycle 9 of RP, the average protein and polysaccharide concentrations of LB-EPS were 14.41 ± 0.79 and 3.38 ± 0.18 mg g1VSS, which were 3.66 and 0.61 mg g1VSS higher than those of RC in cleaning cycle 8, respectively. The PN/PS ratio of RP was 4.26, and it was 0.38 higher than that of RC. In terms of the molecular hydrophobicity, proteins generally

Fig. 5. Interfacial energies profiles between the sludge flocs as a function of separation distance: the LW, AB and EL interfacial energies (a), and the total interfacial energies (b).

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References

Fig. 6. Protein and polysaccharide concentrations of LB-EPSs.

have stronger hydrophobicity than polysaccharides (Jorand et al., 1998), therefore, proteins of LB-EPS are the key factor determining flocs hydrophobicity. The decrease of concentration and PN/ PS ratio of LB-EPS after long-term 2 mg L1 Pb2þ exposure decreased the flocs hydrophobicity, leading the cake layer fouling mitigation. Excessive LB-EPS could weaken the adhesion force between microbial cells, leading to the deflocculation of flocs, thereby intensifying the cake layer fouling (Li and Yang, 2007). In addition, protein in EPS is the primary source of ionizable functional groups, the increase of concentration and PN/PS ratio of LB-EPS would increase the number of positive ions retained in the cake layer, and further intensify the osmotic pressure of the cake layer (Lin et al., 2014). The increasing osmotic pressure of the cake layer would improve the cake layer resistance. Therefore, the increase of concentration and PN/PS ratio of LB-EPS after long-term 6 mg L1 Pb2þ exposure increased the flocs hydrophobicity but decreased the flocs aggregation capacity, leading the cake layer fouling deterioration. 4. Conclusion Long-term effects of Pb2þ on the operating performance and membrane fouling of H/A/O-MBR treating synthetic electroplating wastewater were investigated. 2 mg L1 Pb2þ decreased the cake layer fouling rate but increased the pore blocking rate. 6 mg L1 Pb2þ accelerated the cake layer fouling rate but decreased the pore blocking rate. The increase of low and moderate MW SCFs deteriorated the pore blocking. The flux decline caused by SCFs in RP in cleaning cycle 4 and 9 fitted the standard blocking and cake filtration best, respectively. The increase of high MW SCFs (especially polysaccharides) facilitated the gel layer formation and mitigated the pore blocking. Pb2þ regulated the concentration and PN/PS ratio of LB-EPS, thus changing the flocs hydrophobicity and aggregation capacity, leading the cake layer fouling rate variation. Acknowledgements The authors greatly appreciate the financial support from the State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology (No. 2017DX08), the Shanxi Provincial R&D Program (201603D111019-1) and the Heilongjiang Provincial R&D Program (GJ2017GJ0023). The authors declare that there are no conflicts of interest. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.chemosphere.2019.05.231.

Aftab, B., Khan, S.J., Maqbool, T., Hankins, N.P., 2017. Heavy metals removal by osmotic membrane bioreactor (OMBR) and their effect on sludge properties. Desalination 403, 117e127. APHA, 2005. Standard Methods for the Examination of Water and Wastewater, twentieth ed. American Public Health Association, American Water Works Association and Water Environment Federation, Washington, USA. Chen, R., Nie, Y.L., Hu, Y.S., Miao, R., Utashiro, T., Li, Q., Xu, M.J., Li, Y.Y., 2017. Fouling behaviour of soluble microbial products and extracellular polymeric substances in a submerged anaerobic membrane bioreactor treating low-strength wastewater at room temperature. J. Membr. Sci. 531, 1e9. Chiang, C.W., Ng, D.Q., Lin, Y.P., Chen, P.J., 2016. Dissolved organic matter or salts change the bioavailability processes and toxicity of the nanoscale tetravalent lead corrosion product PbO2 to medaka fish. Environ. Sci. Technol. 50 (20), 11292e11301. Cooper, N.L., Bidwell, J.R., Kumar, A., 2009. Toxicity of copper, lead, and zinc mixtures to Ceriodaphnia dubia and Daphnia carinata. Ecotoxicol. Environ. Saf. 72 (5), 1523e1528. Dutka, B., Kwan, K., 1984. Studies on a synthetic activated sludge toxicity screening procedure with comparison to three microbial toxicity tests. In: Liu, D., Dutka, B.J. (Eds.), Toxicity Screening Procedures Using Bacterial Systems. Marcel Dekker, New York, pp. 125e138. Feng, B., Fang, Z., Hou, J.C., Ma, X., Huang, Y.L., Huang, L.Q., 2013. Effects of heavy metal wastewater on the anoxic/aerobic-membrane bioreactor bioprocess and membrane fouling. Bioresour. Technol. 142 (8), 32e38. Holakoo, L., Nakhla, G., Yanful, E.K., Bassi, A.S., 2006. Chelating properties and molecular weight distribution of soluble microbial products from an aerobic membrane bioreactor. Water Res. 40 (8), 1531e1538. Hong, H.C., Peng, W., Zhang, M.J., Chen, J.R., He, Y.M., Wang, F.Y., Weng, X.X., Yu, H.Y., Lin, H.J., 2013. Thermodynamic analysis of membrane fouling in a submerged membrane bioreactor and its implications. Bioresour. Technol. 146 (10), 7e14. Hou, J., Miao, L.Z., Wang, C., Wang, P.F., Ao, Y.H., Lv, B.W., 2015. Effect of CuO nanoparticles on the production and composition of extracellular polymeric substances and physicochemical stability of activated sludge flocs. Bioresour. Technol. 176, 65e70. Jang, D., Hwang, Y., Shin, H., Lee, W., 2013. Effects of salinity on the characteristics of biomass and membrane fouling in membrane bioreactors. Bioresour. Technol. 141 (4), 50e56. Jang, M.H., Hwang, Y.S., 2018. Effects of functionalized multi-walled carbon nanotubes on toxicity and bioaccumulation of lead in Daphnia magna. PLoS One 13, e0194935. Ji, J., Qiu, J., Wong, F.S., Li, Y., 2008. Enhancement of filterability in MBR achieved by improvement of supernatant and floc characteristics via filter aids addition. Water Res. 42 (14), 3611e3622. Jiang, T., Kennedy, M.D., De, S.V., Nam, S.N., Nopens, I., Vanrolleghem, P.A., Amy, G., 2010. Characterization of soluble microbial products and their fouling impacts in membrane bioreactors. Environ. Sci. Technol. 44 (17), 6642e6648. -Bigne, F., Block, J.C., Urbain, V., 1998. Hydrophobic/hydrophilic Jorand, F., Boue properties of activated sludge exopolymeric substances. Water Sci. Technol. 37 (4), 307e315. Li, C.C., Cabassud, C., Reboul, B., Guigui, C., 2015. Effects of pharmaceutical micropollutants on the membrane fouling of a submerged MBR treating municipal wastewater: case of continuous pollution by carbamazepine. Water Res. 69 (69), 183e194. Li, H., 2016. Research on an MFC-MBR Integrated System for Wastewater Treatment and Membrane Fouling Mitigation. PhD Thesis. Harbin Institute of Technology, Harbin, China. Li, X.Y., Yang, S.F., 2007. Influence of loosely bound extracellular polymeric substances (EPS) on the flocculation, sedimentation and dewaterability of activated sludge. Water Res. 41 (5), 1022e1030. Lin, H.J., Zhang, M.J., Wang, F.Y., Meng, F.G., Liao, B.Q., Hong, H.C., Chen, J.R., Gao, W.J., 2014. A critical review of extracellular polymeric substances (EPSs) in membrane bioreactors: characteristics, roles in membrane fouling and control strategies. J. Membr. Sci. 460 (9), 110e125. Liu, B., Qu, F.S., Liang, H., Gan, Z.D., Yu, H.R., Li, G.B., Bruggen, B.V.D., 2017. Algaeladen water treatment using ultrafiltration: individual and combined fouling effects of cells, debris, extracellular and intracellular organic matter. J. Membr. Sci. 528, 178e186. Ma, J.Y., Quan, X.C., Si, X.R., Wu, Y.C., 2013. Responses of anaerobic granule and flocculent sludge to ceria nanoparticles and toxic mechanisms. Bioresour. Technol. 149 (4), 346e352. MEE, 2016. National List of Hazardous Substances. http://www.mee.gov.cn/gkml/ hbb/bl/201606/t20160621_354852.htm. Rahman, M.L., Sarkar, S.M., Yusoff, M.M., 2016. Efficient removal of heavy metals from electroplating wastewater using polymer ligands. Front. Environ. Sci. Eng. 10 (2), 352e361. Rosenstein, C., 1990. Methane sulfonic acid as an electrolyte for tin, lead and tinlead plating for electronics. Met. Finish. 88 (1), 17e21. Tian, Y., Li, H., Li, L.P., Su, X.Y., Lu, Y.B., Zuo, W., Zhang, J., 2015. In-situ integration of microbial fuel cell with hollow-fiber membrane bioreactor for wastewater treatment and membrane fouling mitigation. Biosens. Bioelectron. 64 (4), 189e195. Wen, Q.X., Wang, Q., Chen, Z.Q., Li, X.Q., Tian, Y., 2016. Effects of Cu on biological

438

Q. Wang et al. / Chemosphere 232 (2019) 430e438

process of wastewater treatment plant (WWTP) in electroplating industrial park. Desalin. Water Treat. 57 (59), 28715e28723. Wen, Q.X., Wang, Q., Li, X.Q., Chen, Z.Q., Tang, Y.C., Zhang, C.J., 2018. Enhanced organics and Cu2þ removal in electroplating wastewater by bioaugmentation. Chemosphere 212, 476e485. Yang, W., Cicek, N., Ilg, J., 2006. State-of-the-art of membrane bioreactors: worldwide research and commercial applications in North America. J. Membr. Sci. 270 (1), 201e211. You, S.J., Tsai, Y.P., Cho, B.C., Chou, Y.H., 2011. Metabolic influence of lead on polyhydroxyalkanoates (PHA) production and phosphate uptake in activated sludge fed with glucose or acetic acid as carbon source. Bioresour. Technol. 102 (17), 8165e8170. Yuan, L., Zhi, W., Liu, Y.S., Karyala, S., Vikesland, P.J., Chen, X., Zhang, H.S., 2015a. Lead toxicity to the performance, viability, and community composition of activated sludge microorganisms. Environ. Sci. Technol. 49 (2), 824e830.

Yuan, Z.H., Yang, X., Hu, A., Yu, C.P., 2015b. Long-term impacts of silver nanoparticles in an anaerobic-anoxic-oxic membrane bioreactor system. Chem. Eng. J. 276, 83e90. Zhang, D.Q., Eng, C.Y., Stuckey, D.C., Zhou, Y., 2017. Effects of ZnO nanoparticle exposure on wastewater treatment and soluble microbial products (SMPs) in an anoxic-aerobic membrane bioreactor. Chemosphere 171, 446e459. Zhang, Y., Zhang, M.J., Wang, F.Y., Hong, H.C., Wang, A.J., Wang, J., Weng, X.X., Lin, H.J., 2014. Membrane fouling in a submerged membrane bioreactor: effect of pH and its implications. Bioresour. Technol. 152 (1), 7e14. Zhu, Y.B., Yu, S.L., Zhang, B., Li, J.F., Zhao, D.S., Gu, Z.Y., Gong, C., Liu, G.C., 2018. Antifouling performance of polytetrafluoroethylene and polyvinylidene fluoride ultrafiltration membranes during alkali/surfactant/polymer flooding wastewater treatment: distinctions and mechanisms. Sci. Total Environ. 642 (15), 988e998.