Applications of membrane bioreactors for water reclamation: Micropollutant removal, mechanisms and perspectives

Accepted Manuscript Review Applications of membrane bioreactors for water reclamation: micropollutant removal, mechanisms and perspectives Jinxing Ma, Ruobin Dai, Mei Chen, Stuart J. Khan, Zhiwei Wang PII: DOI: Reference:

S0960-8524(18)31225-2 https://doi.org/10.1016/j.biortech.2018.08.121 BITE 20404

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Bioresource Technology

Received Date: Revised Date: Accepted Date:

29 June 2018 28 August 2018 29 August 2018

Please cite this article as: Ma, J., Dai, R., Chen, M., Khan, S.J., Wang, Z., Applications of membrane bioreactors for water reclamation: micropollutant removal, mechanisms and perspectives, Bioresource Technology (2018), doi: https://doi.org/10.1016/j.biortech.2018.08.121

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Applications of membrane bioreactors for water reclamation: micropollutant removal, mechanisms and perspectives Jinxing Ma,1,2 Ruobin Dai,1 Mei Chen,1 Stuart J Khan,2 Zhiwei Wang1* 1

State Key Laboratory of Pollution Control and Resource Reuse, Shanghai Institute of

Pollution Control and Ecological Safety, School of Environmental Science and Engineering, Tongji University, 1239 Siping Road, Shanghai 200092, China 2

UNSW Water Research Centre, School of Civil and Environmental Engineering, University

of New South Wales, Sydney, NSW 2052, Australia Email addresses: [email protected] (Jinxing Ma); [email protected] (Ruobin Dai); [email protected] (Mei Chen); [email protected] (Stuart J Khan); [email protected] (Zhiwei Wang)

Bioresource Technology Re-Submitted in August 2018 for MBR Special Issue [VSI: membrane bioreactors]

Corresponding author: *E-mail: [email protected] (ZW Wang).

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Abstract Membrane bioreactors (MBRs) have attracted attention in water reclamation as a result of the recent technical advances and cost reduction in membranes. However, the increasing occurrence of micropollutants in wastewaters has posed new challenges. Therefore, we reviewed the current state of research to identify the outstanding needs in this field. In general, the fate of micropollutants in MBRs relates to sorption, biodegradation and membrane separation processes. Hydrophobic, nonionized micropollutants are favorable in sorption, and the biological degradation shows higher efficiency at relatively long SRTs (3040 days) and HRTs (20-30 h), as a result of co-metabolism, metabolism and/or ion trapping. Although the membrane rejection rates for micropollutants are generally minor, final water quality can be improved via combination with other technologies. This review highlights the challenges and perspectives that should be addressed to facilitate the extended use of MBRs for the removal of micropollutants in water reclamation. Keywords: membrane bioreactor, micropollutant, water reclamation, biodegradation, wastewater treatment

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1. Introduction Limitations to the availability of clean water present major challenges in many parts of the world, increasingly complicated by population growth, urbanization and continuing climate change. For example, 340,000 children under five die every year from diarrhoeal diseases resulting from unsafe drinking water and a lack of access to clean water for hygiene (WHO/UNICEF, 2015). It is reported that over 80% of the world’s wastewater flows back into the ecosystem without being treated (Connor et al., 2017). Meanwhile, there is greater recognition of wastewater as an undervalued and sustainable source of water, energy and nutrients rather than a nuisance (Ma et al., 2013a; Verstraete et al., 2009). Therefore, an understanding of how best to polish and reuse wastewater has gained increasing importance. The appropriateness of various treatment processes for water reuse is dependent upon the intended use of the treated water, and other factors including cost, energy requirements and available physical space for treatment plant. Of the alternatives available, membrane bioreactor (MBR) technology that integrates conventional activated sludge (CAS) process with a physicochemical membrane filtration unit has attracted much attention (Judd, 2010; Krzeminski et al., 2017; Wang et al., 2014b). With the use of microfiltration or ultrafiltration that rejects biomass particles, most of the colloids and part of the solutes, MBRs have distinctive advantages over CAS processes including smaller footprint and sludge production, higher-quality effluent, robustness of operation, and if designed and maintained appropriately, ease of automation (Huang et al., 2010; Ma et al., 2015b; Nguyen et al., 2013). More recent studies show that the MBR performance can be further improved via the introduction of (i) anaerobic processes such as methanogenesis and electrogenesis leading to lower-energy consumption (and even energy recovery) (Ma et al., 2017; Smith et al., 2012) and highly 3

efficient degradation of pollutants (Suneethi & Joseph, 2011), (ii) advanced filtration processes such as nanofiltration (and forward or reverse osmosis filtration) and electrochemical filtration capable of removing smaller-size pollutants (Achilli et al., 2009; Zheng et al., 2017), and (iii) optimized cleaning and operating strategies (Krzeminski et al., 2017). MBRs have been installed and operated in more than 200 countries with a cumulative treatment capacity of the large-scale MBR plants worldwide (the capacity of each plant > 10,000 m3/d) of over 20 million m3/d (Zheng et al., 2016). Continuing global market growth rates of up to 15% have been forecast (Judd, 2016). While enhanced degradation of organic contaminants and removal of nitrogen species have been implicated in MBRs as a result of the unique operating conditions including the long solid retention time, enrichment of specific functional genes/enzymes, etc. (Calderón et al., 2012; Ma et al., 2013b), further consideration should be given to the efficacy of MBRs in treatment of micropollutants (e.g., pharmaceuticals, nanoparticles and antibiotics) (Alturki et al., 2012; Kovalova et al., 2012; Mei et al., 2014). This is particularly important if direct or indirect potable reuse is proposed for the treated water. Micropollutants encompass a vast and expanding array of natural and anthropogenic substances at trace concentrations in water (Gavrilescu et al., 2015; Luo et al., 2014). Compared to the “macropollutants” such as ammonia and polysaccharides, of which the behaviors are relatively well understood in aquatic systems, major challenges remain in understanding the occurrence of micropollutants and their potential health effects as a result of the low concentrations and high diversity. Although there are no legal discharge limits for micropollutants, some regional regulations (European Union Directives 2013/39/EU and 2015/495/EU) have been developed (Barbosa et

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al., 2016). After exhaustive reviews of risks posed by no less than thousands of substances, a number of priority substances/groups of substances such as bisphenol A and the non-steroidal anti-inflammatory, diclofenac, have been selected in watch lists for inter-government monitoring (Barbosa et al., 2016; Luo et al., 2014). Responding to the potential regulations of MBR permeate for water reclamation, intensive studies have been carried out in recent years, towards a better understanding of the behaviors and treatment of micropollutants in MBRs (Chon et al., 2012; Kovalova et al., 2012). However, to date, there has been limited comprehensive summary of the migration and transformation of micropollutants in MBRs as well as the mechanisms underpinning the better removal efficiency compared to the alternatives (e.g., CAS) (Alvarino et al., 2018; Besha et al., 2017). As a result, we review the sorption, biodegradation and membrane separation processes in MBRs that account for the removal of micropollutants. The processes are delineated from the occurrence, transformation and fate of a certain micropollutant in an MBR. 2. Occurrences and fate of micropollutants in MBRs 2.1. Occurrences of micropollutants Previous reviews have well summarized the occurrences and sources of micropollutants in aquatic systems (Luo et al., 2014); for example, industrial, hospital and household effluents containing petrochemicals, pharmaceuticals and personal care products (PCP) are known important point sources for micropollutants in sewages (Barbosa et al., 2016; Gavrilescu et al., 2015). The biodegradation of these products in sewages further generates transformation products such as metabolites of chloroacetanilide herbicide and perfluorinated compounds. Conventional wastewater treatment plants (WWTPs) are not designed to treat

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micropollutants and the removal performances vary significantly among different chemical categories (Deblonde et al., 2011). As such, WWTP effluents have become one main cause of micropollutant contaminations in surface waters (Singhal & Perez-Garcia, 2016). It has been reported that there might be a correlation of pharmaceutical levels present in the effluents of WWTPs with those in freshwater rivers and canals especially in developing countries (Pal et al., 2010), although the attenuation of the concentrations in natural receptors has been observed likely due to water dilution, sorption, direct/indirect photolysis and aerobic biodegradation. With regard to the recent studies of, for example, Alturki et al. (2012), Phan et al. (2015a) and van den Akker et al. (2014), in innovative technologies (such as full-scale MBRs) treating micropollutants in wastewater, Besha et al. (2017) and Verlicchi et al. (2012) reviewed extensively the occurrence and removal of pharmaceuticals in municipal wastewaters. Results indicated that MBRs could be superior to CAS processes in treatment of pharmaceuticals and hormones (Verlicchi et al., 2012) though a previous study concluded that the efficacy of technologies in removing bisphenol-A, ibuprofen and bezafibrate depends on operating conditions (e.g., solid retention time (SRT)) rather than the configurations (either CAS or MBR) (Clara et al., 2005). 2.2. Sorption of micropollutants Generally, the removal mechanisms of micropollutants in MBRs can be classified into (i) sorption, (ii) biological degradation and (iii) membrane separation. Sorption represents the process in which micropollutants become associated with the solid phase. It has been reported that the sorption of micropollutants onto sludge flocs (and bound microbial products) is one of the key factors controlling the removal of micropollutants in wastewater treatment systems

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(Radjenović et al., 2009), while the efficiency largely depends on the physicochemical characteristics of micropollutants (e.g., hydrogen bond, hydrophobicity, electrostatic interactions, etc.) and the sorbents (e.g., mass concentration) (Barret et al., 2010; Mei et al., 2014; Wick et al., 2011). The concentration of a micropollutant adsorbed on activated sludge (Cs, µg/L) can be described by the Freundlich model (Siegrist & Joss, 2012; Wick et al., 2011). Cs in activated sludge can be calculated by the following equation: Cs= Kd·Xss·Caq

(1)

where Caq is the concentration of micropollutant in aqueous phase (µg/L), Xss the concentration of biomass or the amount of sludge produced (g-TSS/L) and Kd the sorption constant (L/g-TSS). It should be noted that this equation assumes that the sorption process is at equilibrium and takes no account of kinetic factors. However, in practice, an MBR system may be far from equilibrium conditions since this can be very slow to be met. Assuming that there is no significant biodegradation, the removal efficiency (ηsorb) of micropollutants in a fully mixed system at equilibrium depends on Xss and Kd as follows: ηsorb = Cs/(Cs + Caq) = Kd·Xss/(1+Kd·Xss)

(2)

In general, Xss is dependent on the concentrations of biomass (and bound microbial products). At this point, the high sludge concentrations in MBRs (up to > 10 g-TSS/L) likely account for their better performance for micropollutant removal compared to CAS processes (Radjenović et al., 2009). With regard to the behavior of micropollutants on the solid phase (i.e., Kd), sorption is expected to be driven by the hydrogen bonding and/or hydrophobic effects with two mechanisms (i.e., adsorption and absorption (or hydrophobic adsorption))

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proposed (Fig. 1a). Adsorption is ascribed to electrostatic interaction of positively charged groups (e.g., quaternary amine) of micropollutants to the negatively charged surface of the cell. The electrostatic interaction is characterized by the acid dissociation constant (pKa) (Sipma et al., 2010). Sorption can also be influenced by intermolecular forces. In the absorption (hydrophobic adsorption) process, micropollutants move from the aqueous phase (i) into the lipophilic cell membrane of microorganisms in activated sludge and/or (ii) hydrophobic surfaces of the sludge matrix. This process can be described by the values of Kow (Carballa et al., 2005), although pKa is also an important consideration because if the chemical is extensively ionized, the hydrophobic interaction will be reduced. Kow is defined as the equilibrium distribution of the ratio of the solubility of a substance in octanol to water. With regard to the speciation of micropollutants at different pH, a more accurate parameter, DpH that simultaneously considers the hydrophobicity and ionization characteristics of micropollutants should be used (Besha et al., 2017). The relationship between DpH and different species can be determined by the following equations:

log DpH = log

[ nonionized + ionized species]octanol [ nonionized + ionized species]water

(3)

The logDpH values can be (i) obtained from the ChemSpider (www.chemspider.com) and OCHEM (www.ochem.eu/home/show.do) databases and/or (ii) calculated using commercial softwares such as ALOGPS program (www.vcclab.org). If only the neutral form participates into the organic phase, the computation of logDpH can be simplified to: For acidic compounds:

(

log DpH = log Kow − log 1 + 10( For basic compounds: 8

pH-pKa )

)

(4)

(

log DpH = log K ow − log 1 + 10(

pK a -pH )

)

(5)

The lipophilicity (logDpH) of micropollutants can significantly influence their sorption tendency to activated sludge. The more hydrophobic the compound is, the much higher sorption tendency and better removal efficiency is normally expected (Alvarino et al., 2018). A number of studies have shown that MBRs exhibit higher removal efficiency for the relatively hydrophobic compounds (logD > 2 - 3) than the hydrophilic (logD < 2 - 3) (Chon et al., 2012; Cirja et al., 2008; Tadkaew et al., 2011). Sahar et al. (2011) reported that sulfonamides (a group of hydrophilic compounds with logDpH=7 < 0.4) were mainly removed by biodegradation with a small fraction immobilized on activated sludge during MBR treatment. Urase et al. (2005) investigated the removal efficiency of 15 micropollutants in MBRs and found that the removal efficiencies of acidic pharmaceuticals were higher in the case of lower pH due to the increasing tendency of sorption to sludge (Eq. 4). It should be noted that the hydrophobic interaction may be also expected for the compounds in the form of zwitterion having low logDpH (e.g., oxytetracycline) (Cirja et al., 2008). In summary of the two mechanisms contributing to the sorption of micropollutants (Fig. 1a), Kd can be derived from Kow values for non-polar compounds. In contrast, Kd values of polar and charged compounds are always evaluated by LogDpH. It can be seen from Fig. 1b that Kd values of micropollutants are generally well correlated with their LogDpH values as Kd is a measure of adsorption and LogD is a strong predictor for hydrophobic adsorption. Specifically, the removal of most analgesics and anti-inflammatories by sorption should be a minor pathway in MBRs as their Kd < 0.5 L/g-TSS (Ternes et al., 2004) (Fig. 1b). For those with Kd > 5 L/g-TSS (such as hormones and endocrine disrupter compounds and musk

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fragances), MBR treatment is capable of improving effluent quality via the immobilization of these compounds on the activated sludge, though the complete detoxification relies on the biological degradation. Moreover, the determination of apparent Kd values can be interfered by the coexisting processes (e.g., biological degradation). For example, no sorption onto the sludge should be expected for acetaminophen while Radjenović et al. (2009) obtained an extremely high value for Kd of acetaminophen for secondary activated sludge (1.2 L/g-TSS). A plausible explanation relates to a very fast biodegradation that outcompeted the sorption process and led to very low concentrations measured in the aqueous phase (Radjenović et al., 2009). 2.3. Biological degradation Biological degradation of micropollutants is an essential aspect of MBR treatment. Otherwise, the continuous sorption onto the sludge will lead to either the exhaustion of the capacity and deterioration in the performance or enrichment of micropollutants in the solid phase that may cause secondary pollution. As the biodegradability of micropollutants relates to their bioavailability, the compound structure/property is also involved in determining the resistance of a micropollutant to biodegradation other than sorption abovementioned. For example, Tadkaew et al. (2011) proposed a qualitative framework to predict the removal of trace organic micropollutants in MBRs based on their logDpH values (Eqs. 3-5). Low removal efficiency (< 20%) was implicated for hydrophilic compounds (logDpH=8 < 3.2) with strong electron withdrawing groups (e.g., nitro and amides) while biodegradation becomes efficient when strong electron donating groups are dominant in the hydrophilic compounds (Tadkaew et al., 2011).

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2.3.1. Aerobic degradation Aerobic biodegradation of micropollutants is widely implemented in MBRs, which exploits the catabolic activities of microorganisms that assimilate target micropollutants as growth substrates followed by their elimination (Tran et al., 2013). The growth condition of aerobic microorganisms is closely related to the parameters such as SRT. It has been accepted that the higher biomass concentration in MBRs resulting from the long-SRT operation is favorable in the biotransformation of some micropollutants compared to CAS processes (Bernhard et al., 2006; Radjenović et al., 2009). Meanwhile, long SRTs allow an enrichment of slowly growing autotrophic bacteria such as nitrifiers or others that may benefit the biodegradation of micropollutants. A more recent study has indicated that a longer SRT can lead to higher rRNA expressions for rare taxa as well as diverse rRNA expression patterns between redox periods (Vuono et al., 2016), which may explain why MBRs with longer SRTs always have higher rates of micropollutant biotransformation. For standardization, the pseudo-first order degradation kinetics is determined as follows (Eq. 6)(Gschwend, 2016):

dC = Kbiol X ssCaq dt

(6)

where C is the total concentration of the micropollutant (µg/L), t the reaction time (d), Kbiol the pseudo first order reaction rate constant (L/(g-TSS d)) and Caq the concentration of the soluble part of the compound (or in aqueous phase, µg/L). Fig. 2a summarizes the Kbiol values for selected micropollutants in MBRs (and CAS processes for comparison). In agreement with the qualitative framework developed by Tadkaew et al. (2011), the biological treatment shows low removal efficiency (Kbiol < 1 L/(g-TSS d)) for the organic compounds with strong electron withdrawing groups (Besha et al., 2017). While MBRs overall have higher oxidation 11

capacity than the CAS processes, unexpected high Kbiol values were observed in lab-scale conventional activated sludge reactors, working under nitrifying and denitrifying conditions (e.g., the stars in Fig. 2a) (Suarez et al., 2010). Micropollutants are present in wastewaters at trace concentration levels ranging from ng/L to µg/L (Kolpin et al., 2002; Stasinakis, 2012), and many of them can cause toxicity and/or resistance to microbial growth. Therefore, it is expected that the energy generated from biodegradation of micropollutants is not sufficient to support the microbial growth and that primary substrates are necessary to initiate the biotransformation. As a result, in either an MBR or CAS system, co-metabolism is regarded as the main pathway in biodegradation of micropollutants (Fig. 2b) (Alvarino et al., 2018). Co-metabolism has been intensively studied in the nitrification process that commonly exists in aerated MBRs (and CAS processes) (Mazioti et al., 2015). For example, it is reported that a nitrifier enrichment culture from an MBR can biotransform 17αethinylestradiol (EE2) with ammonia oxidation as the driving force (Gusseme et al., 2009). A maximum EE2 removal rate of 9.0 µg/(g-VSS h) was achieved with >94% removal efficiency. In another study of MBRs (Maeng et al., 2013), there was a positive relationship between the biotransformation of acidic pharmaceuticals (i.e., gemfibrozil, diclofenac, bezafibrate and ketoprofen) and a synthetic estrogen, EE2, and the nitrification process. The ammonium monooxygenase (AMO) is considered as the key enzyme responsible for the cometabolism of micropollutants/NH3 (Fig. 2b) because co-metabolic biotransformation of estrogens and pharmaceuticals can occur in the isolates of autotrophic ammonia-oxidizing bacteria (AOB) (Khunjar et al., 2011; Roh et al., 2009; Tran et al., 2013). Although the use of

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membranes in MBRs may cause discrepancy in microbial communities compared to CAS processes treating the same micropollutants, the mechanisms underpinning the biotransformation should be similar. Helbling et al. (2012) found that in addition to the bacterial amoA, the archaeal amoA gene expression was noticed as well in the sludge communities during the biotransformation of micropollutants. The transcript abundance of the archaeal amoA has a positive correlation to the biotransformation of isoproturon, venlafaxine and ranitidine. Such conclusion may assist in interpreting the outcomes of the study of Kruglova et al. (2017), in which the MBR treating ibuprofen, diclofenac, estrone and EE2 exhibited high nitrifying efficiency (> 98%) and comparable micropollutant removal efficiency but having low abundance of AOB (< 0.5%) and heterotrophic nitrifies. In regard to the enrichment of nitrifiers and amo in MBRs (Ma et al., 2016), one can expect their higher oxidation capacity compared to CAS processes (Fig. 2a). Moreover, other enzymes involving wastewater treatment can be functionally versatile. For example, Men et al. (2016a) reported that AOB were highly involved in the degradation of asulam, clomazone, monuron and trimethoprim, while ATU (allylthiourea)-affected enzymes (likely coppercontaining enzymes not critical for heterotrophic respiration) other than AMO were even more important for certain micropollutants. According to the studies of Fernandez-Fontaina et al. (2016), Nguyen et al. (2016) and Zhang et al. (2004), consideration should also be given to the roles of phenol hydroxylase, esterase, phthalate dioxygenase and laccase playing in micropollutant biotransformation in MBRs. Although less studied than co-metabolism, the heterotrophic pathway (metabolism) for micropollutant biotransformation under aerobic conditions has been reported recently (Fig.

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2b) (Fernandez-Fontaina et al., 2016). A pure culture Sphingomonas Ibu-2 is capable of degrading ibuprofen as the sole carbon source (Murdoch & Hay, 2005), while the heterotrophic culture Delftia tsuruhatensis and P. aeruginosa can utilize acetaminophen as the sole carbon source in an MBR (De et al., 2011). Zhang et al. (2013) found that three isolated strains (Stenotrophomonas sp. strain f1, Pseudomonas sp. strains f2 and fg-2) could grow on paracetamol as the sole carbon, nitrogen, and energy sources and that the combination of them led to significant improvement in degradation and mineralization of paracetamol. The biotransformation of sulfamethoxazole may also relate to the heterotrophic metabolism rather than the nitrification process (Alvarino et al., 2016). So far, micropollutant biotransformation via metabolism process has received much less attention compared to cometabolism, and whether the heterotrophic pathway plays a vital role in MBRs treating micropollutants is open to further investigation. In addition to the co-metabolism and metabolism mechanisms, a novel theory relating to micropollutant removal has been proposed. It is reported that the protozoa, which make up about 5-10% of the activated sludge biomass (Pauli, 2013; Ratsak et al., 1996), can remove aliphatic amines that are highly abundant among organic micropollutants from the aqueous phase, through ion trapping (Gulde et al., 2018). In ion trapping, the neutral species of aliphatic amines diffuse across the cell membrane and further enter acidic vesicles in eukaryotic cells, such as protozoa. Then the amines are trapped since diffusion of positively charged species formed in the acidic vesicles is strongly hindered (Gulde et al., 2018) (Fig. 2b). Notably, a specific type of acidic vesicles, e.g., acidocalcisomes, are also present in bacteria, particularly in phosphate-accumulating bacteria (Docampo & Moreno, 2011). These

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findings indicate that the investigation on protozoa and phosphate-accumulating bacteria treating micropollutants should be considered in future studies of MBRs. 2.3.2. Anaerobic degradation Different from aerobic biodegradation, anaerobic biodegradation occurs in the absence of bound and dissolved oxygen. Anaerobic membrane bioreactor (AnMBR) that combines anaerobic and membrane technology is a sustainable alternative for low-strength wastewater treatment (Ma et al., 2013c). Compared to the aerobic condition, one may expect that the removal of at least a small group of micropollutants that present limited aerobic biodegradation potential, such as diatrizoate (Redeker et al., 2014), triclosan (Gangadharan et al., 2012), trimethoprim (Alvarino et al., 2014), venlafaxine (Falås et al., 2016), carbamazepine (König et al., 2016) and sulfamethoxazole (Falås et al., 2016) can be enhanced by anaerobic operation of MBRs. In addition, the combination of anaerobic and aerobic processes is allegedly capable of improving the removal of specific and non-specific in vitro toxicities of 31 common micropollutants and 10 metabolites (Völker et al., 2017). So far, there has been limited investigation on the relationship between the anaerobic populations (and their functions) and the removal of micropollutants. Harb et al. (2016) found that in an AnMBR fed with synthetic wastewater containing 27 kinds of micropollutants, the dominant genes bphA1, xylA and glx showed strongest correlation to micropollutant biodegradation. Additionally, the long chain fatty acid CoA ligase, which is known to target the carboxyl or hydroxyl groups as an initial step of some micropollutant conversion, was proven to be important for biotransformation in AnMBRs (Harb et al., 2016). Gonzalezgil et al. (2017) demonstrated that a key enzyme in methanogenesis, acetate kinase, can transform

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galaxolide, naproxen, nonylphenol, octylphenol, ibuprofen, diclofenac, bisphenol A, and triclosan. They also found that the action of acetate kinase was chemical structure-dependent to micropollutants, which means that only compounds that contain a carboxyl or hydroxyl group and have moderate steric hindrance are transformable. In addition, the action of acetate kinase accounted for 10 - 90% of the methanogenic biotransformation of micropollutants (Gonzalezgil et al., 2017), indicating that substantial studies are required to understand the roles of other active enzymes playing in the anaerobic biotransformation. 2.4. Membrane separation processes Membrane separation is another mechanism contributing to compound removal in MBRs, as a result of size exclusion and charge repulsion (Wang et al., 2013a). Direct sorption of micropollutants onto the membrane surface may not be important in microfiltration (MF)/ultrafiltration (UF) MBRs because the molecular sizes of micropollutants are generally smaller than the pore sizes of MF/UF membranes (Sahar et al., 2011). However, the inefficiency can be overcome by using fine-pore membranes (e.g., nanofiltration (NF)/reverse osmosis (RO) filtration membranes). Moreover, it has been reported that the removal of micropollutants may also been achieved in MF/UF MBRs due to the rejection/sorption by the foulants that act as a secondary barrier on membranes (Li et al., 2011; Urase et al., 2005). As discussed in literature (Comerton et al., 2007; Sahar et al., 2011), the removal of relative nonpolar substances by the sorption of membrane occurs due to the hydrophobic interaction although this effect is not important for the hydrophilic compounds. Schröder (2002) indicated that polar compounds could be partially removed by MF in the activated sludge process, while Reemtsma et al. (2002) reported that the contribution from MF removal

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was negligible. Yoon et al. (2006) investigated the rejection performance of UF and NF for 52 kinds of endocrine disrupting compounds, pharmaceuticals and personal care products under non-equilibrium conditions. High removal efficiency was observed for nonpolar and hydrophobic compounds (Yoon et al., 2006). According to Alvarino et al. (2017), the removal efficiency of diclofenac and rocithromycin in an ultrafiltration MBR was higher than that in a microfiltration MBR due to the retention by the cake layer. Studies of micropollutant removal by the membrane separation process demonstrate that the sorption of compounds not only occurs on the membrane surface but in the membrane skin layers, the pores and the support layers (Chon et al., 2012) though it is challenging to distinguish the contributions of different parts of membranes to sorption. Regarding to the size exclusion effects, it can be expected that the efficacy of membrane separation is related to the pore size (Yoon et al., 2004). Nevertheless, fine pores may affect the interaction of micropollutants with the different layers of membranes; for example, Comerton et al. (2007) evaluated the sorption of micropollutants by UF/NF/RO membranes and the results showed that larger pore sizes could allow more access to the membrane’s internal adsorption sites, thus increasing the sorption of compounds. 3. Factors affecting the process performance 3.1. Types of membranes and MBRs for micropollutant removal Generally, MBR configurations can be divided into two types, i.e., side-stream (external or recirculated) MBR and submerged (immersed or integrated) MBR (Wang et al., 2008; Wang et al., 2013c). In a side-stream MBR, the membrane module is located outside the bioreactor. With regard to the high energy cost of the recirculation pump in the side-stream

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configuration, the concept of submerged MBRs has been proposed by immersing membrane modules in an aerated tank that allows the effluent to pass through while retaining the particles. Considering the mechanisms (e.g., sorption, biodegradation and MF/UF membrane separation) involved in the micropollutant removal, one should expect there is little difference in the efficiency between these two configurations though the use of submerged MBRs can decrease the operation cost (Besha et al., 2017). The introduction of fine-pore membranes such as NF/RO membranes in submerged MBRs can improve the performance (Tay et al., 2018). However, the high biomass concentrations in MBRs can readily lead to the formation of cake layers during high-pressure filtration (Meng et al., 2009; Meng et al., 2017). As a result, a two-stage membrane filtration process has been proposed and implemented (by, such as Origin Water Co., China) to polish the MBR permeate; that is, a side-stream NF/RO unit is subsequently set up to treat the permeate from a submerged MBR (or a side-stream MBR (Tay et al., 2018)), and the NF/RO retentate would be recirculated to the MBR unit to avoid (i) the accumulation of micropollutants in the NF/RO unit and (ii) severe membrane fouling occurring in the co-presence of organics and dissolved silica (Kimura et al., 2016). 3.2. Operating conditions and principles 3.2.1. SRT, HRT and F/M ratio SRT (i.e., solid/sludge retention time) and HRT (i.e., hydraulic retention time) are two of the most important parameters for MBR design and operation (Judd, 2010). For a given kind of wastewater, SRT and HRT can also influence other parameters such as organic loading rate and food to microorganism ratio (F/M ratio). While the individual management of SRT and

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HRT is one of the advantages of MBRs over CAS process thus allowing the conservation of long generation-period bacteria, it should be noted that complete decoupling of SRT and HRT cannot be achieved in practical applications because the unrestrained extension of SRT while decreasing the HRT would result in very high biomass concentrations (and low F/M ratio). The consequent microbial endogenous respiration and decomposition of the biomass can cause severe membrane fouling and deterioration of the effluent quality (Wang et al., 2013b). Figs.3a and b show two studies on the influence of SRT and HRT on MBR performance treating micropollutants. It can be seen that MBRs have a high removal efficiency (>80 90%) for a group of compounds including acetaminophen, ibuprofen, bezafibrate, E1, E2, etc., even at a very low SRT = 8 d (Maeng et al., 2013). This observation is in good relationship with the biological degradation kinetics of these micropollutants as their Kbiol values are generally over 1~10 L/(g-TSS d) (Fig. 2a). In contrast, carbamazepine and diclofenac are hard to remove from the effluents regardless of the SRT used (Fig. 3a). This is not surprising because these hydrophilic compounds (logDpH=7-8 < 3.2) with strong electron withdrawing groups are difficult to decompose in biological processes (Besha et al., 2017; Tadkaew et al., 2011). Of particular interest is to notice that the biological degradation of gemfibrozil, EE2 and naproxen (Kbiol = ~1 L/(g-TSS d)) depends on the SRT and an increase from 8 to 80 d can significantly improve the efficiency (Fig. 3a). This may suggest that the optimization of the SRT of MBRs treating micropollutants should be focused on this group of compounds. Monitoring the change of the microbial community structure (such as via highthroughput sequencing) (Ma et al., 2015a) can provide insights into the strains that are responsible for the removal of, for example, gemfibrozil. The understanding of their

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proliferation kinetics would be useful to determine the optimum SRTs. According to the literature studies (Fig. 3c), an SRT of 30 - 40 d can be recommended. HRT stands for hydraulic retention time, referring to the average retention time of any molecule of water in an MBR. Overall, it can be expected that an extension of HRT would lead to a higher removal rates of the micropollutants (Fig. 3b), although some studies reported that the impacts of HRT in the range from 7 to 14 h on micropollutant removal may be not significant (Weiss & Reemtsma, 2008). Determination of the optimum HRT in MBRs also relates to the consideration of capital and operating costs. For example, lager footprint of the bioreactor is required at high HRTs. Therefore, an HRT between 20 and 30 h is widely used in MBRs treating diverse micropollutants (Fig. 3c). With regard to the SRTs and HRTs implemented, the F/M ratio in these MBRs is around 0.15 kgCOD/(kgMLSS d). 3.2.2. pH and temperature While the pH values of most municipal wastewater are within the circumneutral range, the shock load due to industrial streams may acidize (or alkalify) the wastewater. pH can affect both sorption and biological processes. A number of analgesics and anti-inflammatories (e.g., ibuprofen, diclofenac and ketoprofen) are ionisable pH dependent. It has been reported that MBRs have higher removal efficiencies for these compounds at low pH conditions (Sanguanpak et al., 2015; Tadkaew et al., 2010), largely due to the increase in the lipophilicity (Eqs. 3-5). Nevertheless, the optimum pH for the biological process in MBRs is around 6 - 7, and the microbial activity for biodegradation and nitrogen metabolism would decrease at low pH. Moreover, Sanguanpak et al. (2015) investigated the micropollutant removal and membrane fouling in MBR treatment of landfill leachate at various pH levels

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and found that pH 5.5 led to the most severe membrane fouling due to the deposition of proteins and carbohydrates. As such, the removal of ionisable species can be achieved by alternatives such as the addition of adsorbent materials rather than change of the solution pH. It is accepted that there is an apparent relationship between the ambient temperature and the sorption and degradation of organic compounds. As for the sorption process, the Gibbs free energy change (ΔG), enthalpy change (∆H) and entropy change (∆S) are three important parameters used to reflect the thermodynamics. The relationship between ΔG, ΔH and ΔS can be expressed using the following Eqs.7 and 8 at equilibrium: ln Caq =

∆H RT

∆S   +  ln Q −  R  

(7)

ΔG =ΔH – TΔS

(8)

where R is the universal gas constant (8.314 J/(K mol)), T the temperature (K) and Qe the amount of adsorption at equilibrium (mg/g) that may be simplified to Cs/Xss at steady state. ∆H can be used to predict whether the sorption process is an endothermic/exothermic type. As for the effects on the biological process, an increase in temperature in an appropriate range is conducive to micropollutant removal. For example, Suárez et al. (2012) evaluated the removal efficiency of sulfamethoxazole and erythromycin as a function of temperature and the results indicated that an increase in temperature resulted in 30% removal than that in cold weathers. However, increasing temperature to a high level (e.g., 45 ℃) can inhibit the metabolic activity (Besha et al., 2017). Likewise, low temperature also deteriorates the treatment efficiency. Attention should be given to the seasonal change of the performance especially in frigid zones. For example, Gurung et al. (2017) assessed the performance of an MBR plant to treat real municipal wastewater during winter season in 21

Nordic regions. They found the elimination of diclofenac was higher in Jan-Feb than MarchApril because the higher liquor temperature in Jan-Feb (15 ℃) than March (<10 ℃) (Gurung et al., 2017). The boundary temperature should therefore be taken into account in practical applications and emergency measures (such as change of SRT/HRT and thermal insulation) can be applied in the extreme conditions. 3.2.3. Oxidation-Reduction Potential (ORP) According to the presence/absence of difference oxygen species, the MBR conditions can be classified into aerobic (i.e., respiration in the presence of free oxygen, ORP > 50 to 100 mV), anoxic (i.e., in the absence of free oxygen but the presence of bound oxygen, ORP ranging from 100 to -100 mV) and anaerobic conditions (i.e., cellular respiration in the absence of oxygen, ORP < -100 to -50 mV). MBRs can organize different microbial diversity and enzymatic functions and activities at different ORPs (Judd, 2010). It has been reported that musks (galaxolide, tonalide and celestolide) and estrogens (E1 and E2) can be well degraded under both aerobic and anoxic conditions (Suarez et al., 2010). Transformation of ibuprofen, EE2, roxithromycin, erythromycin, citalopram and naproxen is only expected in the aerobic process. In contrast, the degradation of diclofenac, sulfamethoxazole, diazepam, trimethoprim and carbamazepine is much less efficient in the presence of oxygen species (Suárez et al., 2012). Hydrogenation of these compounds with strong electron withdrawing groups was observed under anaerobic conditions (e.g., carbamazepine to 10,11-Dihydrocarbamazepine) (König et al., 2016). However, the reasons for these variations in the removal efficiency as a function of ORP are still not fully understood and therefore further studies regarding the relevance of micropollutant transformation and microbial metabolisms under

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different ORP conditions are necessary. 3.2.4. Pilot-scale/full-scale studies While lab-scale studies can provide important information on (i) understanding the mechanisms involving the removal of micropollutants in MBRs and (ii) optimizing the operating conditions to improve the treatment efficiency, the laboratory conditions might be unable to mimic practical conditions well enough, particularly when synthetic feed is used. Table 1 summarizes four case studies on micropollutant removal in pilot-scale/full-scale MBRs treating municipal wastewater. Under similar operating conditions (compared to Fig. 3), the MBR systems can generally achieve high micropollutant removal efficiency with the effluent quality largely complying with the local guidelines for water recycling. These results indicated that lab-scale studies on the factors affecting the process performance could still provide insight to scale-up of the technology for applications. Moreover, during the pilotscale/full-scale operation of MBRs, consideration should be given to the impacts of hazardous events on the treatment performance. For example, the study on aeration failure, power loss, and chemical shocks (ammonia or bleach) showed that the removal of hydrophilic micropollutants that are resistant and/or occur at high concentrations could be significantly affected by the accidents (Phan et al.,2015a), with chemical shocks temporarily increasing the endocrine activity of the effluent. 4. Challenges and perspectives 4.1. Challenges in MBRs treating micropollutants 4.1.1. Identification of micropollutants and understanding of their health risks While there have been progress and developments in high performance liquid

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chromatography and/or gas chromatography mass spectrometry in last decades that greatly facilitates the environmental research on emerging polar/nonpolar organic pollutants (Hermes et al., 2018), only a small fraction of the total number of potential contaminants are taken into consideration in recent studies. For example, nano-particles released or transformed from nano-materials can interact with the biosphere via numerous ecological processes (Zheng et al., 2011). While the adverse effects of the “nano-particles” on MBR performance have been emphasized in the studies of Wang et al. (2014a) and Tan et al. (2015), the measurement and characterization of nano-particles (such as ZnO) and their hydrolysis products in wastewaters and MBRs are not without challenges. Also, there is very limited knowledge on the toxicology of pollutants at extremely low levels in the environment although these compounds may act in the same way intended for their application such as neurotoxic insecticides damaging the nervous systems of aquatic organisms (Thomaidi et al., 2015). Considering that MBRs can be used as a nexus of waste, water, resource and energy, more research is required to identify the sources and presence of recalcitrant micropollutants in wastewaters and to elucidate the acute and chronic toxic effects. The potential consensus may include an updated list of priority pollutants ratified and implemented over governments. 4.1.2. Impacts of micropollutants on MBR microbial activities and membrane filtration As discussed in Section 2.2., sorption can be an important contribution to micropollutant removal in MBRs. However, it should be noted that this is a transfer process rather than a transformation process. The micropollutants can be enriched in activated sludge (Alvarino et al., 2018; Urase & Kikuta, 2005). On condition that these micropollutants cannot be degraded in the subsequent biological processes (Fig. 1), attention should be given to the treatment of

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waste sludge as improper disposal would cause secondary pollution (Wang et al., 2013b). Hydrophobic interaction (and absorption) that relates to the migration of organics from the aqueous phase into the lipophilic cell membrane of microorganisms has been identified as an important mechanism contributing to micropollutant removal in MBRs. However, it has been reported that the primary mechanisms for the antimicrobial agents also involve in bacterial disorganization that commences with the binding of the agents to cell membranes by ionic and/or hydrophobic interactions (Banerjee et al., 2011). Our recent studies showed that a group of cationic surfactants, quaternary ammonium compounds (QACs), might disrupt the integrity of the cell walls (and membranes) as well as cause induced feedback (e.g., inhibition of respiratory enzymes and dissipation of the proton motive force) (Zhang et al., 2016). While the inhibition effects might be offset by the biomass at high concentrations, consideration should be given to the chronic impacts on the long-term operation of MBRs. Fig. 4 depicts a comprehensive procedure to evaluate the interaction of micropollutants with microorganisms (e.g., absorption) and its influences on MBR performance. The process shown in Fig. 4 begins with the evaluation of the acute toxicity of selected compounds on microorganisms. An increase in the concentrations of microbial products with no significant change in their adhesion and fluidity properties may indicate relatively low adverse impacts on microbial activities. In contrast, an increase in oxidation stress, loss of metabolism activities and/or change in integrity relate to severe cell damage (Fig. 4). The variation of the acute response as a function of time (or chronic toxicity) should subsequently be evaluated in an MBR under continuous operation (Zhang et al., 2018). The direct deposition of extracellular polymeric substances (EPS) on the membrane

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surface and/or indirect interaction (i.e., dissociation of EPS that generates soluble components termed soluble microbial products (SMP)) can cause severe membrane fouling in MBRs (Meng et al., 2009; Wang et al., 2009). It has been reported that the addition of cyclophosphamide and its principal metabolites (Avella et al., 2010; Delgado et al., 2010), carbamazepine, diclofenac, ibuprofen and naproxen (Lay et al., 2012; Li et al., 2015) at environmental concentrations (µg/L) can induce an increase in the bound (and soluble) EPS. Specifically, the EPS formed on exposure to ~5 µg/L cyclophosphamide and its principal metabolites were polysaccharides of about 6 kDa (Avella et al., 2010). In contrast, an increase in the protein/polysaccharide ratio in EPS was observed following the addition of carbamazepine, diclofenac, ibuprofen and naproxen in an MBR (Lay et al., 2012). Generally, cell death and hydrolysis should be a reason for the increase in protein/polysaccharide ratio in the bioreactor. The evaluation procedure shown in Fig. 4 is capable of capturing the microbial response to the presence of micropollutants, which can elucidate the biotransformation of cellular substances into the bulk solution. The use of a quartz crystal microbalance with dissipation monitoring (QCM-D), for example, can largely describe the variation of EPS components and their fouling propensities as an interpretation of the ∆D/∆f from QCM-D results can provide insights into the viscoelastic properties of the components during deposition (Chen et al., 2017). Moreover, the addition of micropollutants may change the floc size of the sludge although Delgado et al. (2010) reported that the floc size of the sludge increased on exposure to 5 µg/L cyclophosphamide and its principal metabolites while deflocculation was induced by the presence of carbamazepine (Aubenneau et al., 2010). Small floc size contributes more

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to fouling than larger ones (Le-Clech et al., 2006; Wang et al., 2014b). The complexation of EPS and/or SMP with the micropollutants (e.g., trace metal ions) may also aggravate the membrane fouling; for instance, our recent study showed that the escalation of fouling in ZnO nano-particle presence could also be due to the interaction of ZnO with SMP that formed a dense layer on the membrane surface (Mei et al., 2014). 4.2. Future perspectives In order to facilitate the application of MBRs in water reclamation, consideration should be given to the following aspects when treating micropollutants.


Micropollutant and metabolite monitoring. With an increasing demand for the chemical industrial products worldwide, there can be higher occurrence of diverse micropollutants in wastewaters. Fast, accurate and on-line detection and monitoring methods are in great need of. Helbling et al. (2010) developed an efficient procedure to allow for high-throughput elucidation of transformation product structures via the combination of HPLC and linear ion trap-orbitrap mass spectrometry. Moreover, the evaluation of micropollutant removal in MBRs should be from a lifetime view. Complete mineralization is difficult to achieve in bioreactors. Consideration must be given to the formation of transformation products because it has been reported that a few pharmaceutical metabolites can be as toxic as or more toxic than their parent compounds (Helbling et al., 2010).




Enhanced biotransformation of organic compounds with strong electron withdrawing groups. As discussed in Sections 2.3 and 3.2, MBRs have very limited removal efficiency for organic compounds with strong electron withdrawing groups.

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Recent studies indicated that the biotransformation of this group of micropollutants might require specific conditions (Alvarino et al., 2014; Falås et al., 2016; König et al., 2016). While there has been a number of studies relating the MBR microbial assembly and operating parameters to the high performance over CAS processes under aerobic and/or anoxic conditions (Ma et al., 2016; Radjenović et al., 2009), investigation on how best to implement MBRs, for example, under the anaerobic condition for dehalogenation/hydrogenation, is relatively limited. Further studies to improve the biotransformation of the refractory micropollutants are necessary.


Integrated MBRs to alleviate microbial inactivation. Some micropollutants may have high Kd values but lower Kbiol values, which likely results in their accumulation in the sludge phase. There have been critical concerns about (i) the acute and chronic toxicity of the residue to the microorganisms, (ii) performance deterioration especially in longterm operation and (iii) generation of secondary pollution. The use of absorbent materials such as activated carbon is one of the alternatives available (Alvarino et al., 2017; Li et al., 2011), and membrane separation in MBRs can efficiently prevent the washout of absorbents. In addition, recent studies have shown that the application of specific enzymes such as laccase can also reduce the risk of developing bacteria resistant to chemicals (Nguyen et al., 2016).




Innovative membrane materials and filtration process. Currently, MBRs cannot serve as absolute barriers to micropollutants, partially as a result of the inefficiency of microfiltration/ultrafiltration processes. The application of nanofiltration and forward or reverse osmosis filtration can significantly improve the treatment performance (Dolar et

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al., 2012; Tay et al., 2018), although consideration should be given to the relevant increase in the energy consumption. Another challenge for the use of high-pressure membranes relates to the generation of concentrated stream. As a result, the coupling of membrane separation with the advanced oxidation process (AOP) has recently been proposed (Zaky & Chaplin, 2014). It has been reported that the redox reactions occurring on the conductive membrane surface (under either anodic or cathodic polarization) can (i) facilitate the transformation of refractory pollutants that is an obvious advantage over CAS processes, (ii) alleviate membrane fouling whilst (iii) cause negligible effects on the microbial activities in the bulk solution (Huang et al., 2015; Zheng et al., 2017; Zheng et al., 2018). Future studies may include the improvement of the selectivity of the oxidants towards targeted micropollutants. 5. Conclusions MBR technology is an attractive alternative to overcoming the challenges raised by micropollutants during water reclamation. The fate of micropollutants in MBRs is associated with sorption, biodegradation and membrane separation processes. The sorption efficiency depends on the hydrophobicity and ionization characteristics of micropollutants, while biological degradation (co-metabolism, metabolism and ion trapping mechanisms) shows higher efficiency due to the high biomass concentrations and long SRTs (30-40 days) in MBRs compared to CAS processes. The sustainable application of MBRs requires further understanding of the fate of micropollutants, investigation on the biotransformation mechanisms and combination of MBRs with emerging technologies. Acknowledgement

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This work was financially supported by National Natural Science Foundation of China (Grant 51838009). Dr. Jinxing Ma acknowledges the receipt of a UNSW Vice-Chancellor’s Postdoctoral Research Fellowship (RG152482). Declarations of interest: none Appendix A. Supplementary data

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Figure captions Fig. 1. (a) Sorption mechanisms accounting for micropollutant removal in an MBR and (b) change of log (Kd × 10 3) as a function of logDpH. Data in Fig. 1b (at pH = 7.4) were calculated based on Supplementary Data via ALOGPS program. Fig. 2. (a) Ranges of degradation rate constants, Kbiol, for selected micropollutants in MBRs and CAS processes. Data are retrieved from Abegglen et al. (2009); Besha et al. (2017); Fernandez-Fontaina et al. (2013); Joss et al. (2004); Joss et al. (2006); Suárez et al. (2012); Suarez et al. (2010). The relatively high values (stars in Fig. 2a) are obtained from Suarez et al. (2010). The dashline represents a somewhat arbritrary boundary between slowly (Kbiol < 1 L/(g-TSS d)) and easily (Kbiol > 1 L/(g-TSS d)) biodegradable micropollutants. (b) Biological pathways involved in micropollutant transformation in MBRs. Fig. 3. Micropollutant removal efficiency as a function of (a) SRT and (b) HRT. (c) Typical SRT, HRT and F/M ratio used for MBRs treating micropollutants. In Fig. 3b the C/N ratio stands for the carbon to nitrogen ratio of the influent wastewater. Data in Figs. 3a and b are collected from Maeng et al. (2013) and Boonnorat et al. (2016) respectively. Data in Fig. 3c include the studies by Besha et al. (2017), Boonnorat et al. (2016), Delgado et al. (2011), Gurung et al. (2017), Harb et al. (2016), Kruglova et al. (2016), Sanguanpak et al. (2015) and Zolfaghari et al. (2015). Fig. 4. A comprehensive procedure evaluating the interaction of micropollutants with microorganisms and the influences on MBR performance. The left module represents the exposure procedure to determine the acute toxicity of the selected compound, the middle module illustrates the responding criteria (Zhang et al., 2018), and the right module depicts a

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standard MBR setup to investigate the chronic effects on the process performance. ROS is the acronyms of reactive oxygen species. Previous studies have described how to determine the microbial products (Han et al., 2013), how to characterize the adhesion and fluidity properties (Mei et al., 2014), how to determine the biodegradation and nitrogen metabolism kinetics (Han et al., 2016; Song et al., 2018), and how to evaluate the long-term performance in a flow system (Wang et al., 2014a).

Table captions Table 1. Four case studies on micropollutant removal in pilot-scale/full-scale MBRs treating municipal wastewater.

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Fig. 1. (a) Sorption mechanisms accounting for micropollutant removal in an MBR and (b) change of log (Kd × 10 3) as a function of logDpH. Data in Fig. 1b (at pH = 7.4) were calculated based on Supplementary Data via ALOGPS program.

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Fig. 2. (a) Ranges of degradation rate constants, Kbiol, for selected micropollutants in MBRs and CAS processes. Data are retrieved from Abegglen et al. (2009); Besha et al. (2017); Fernandez-Fontaina et al. (2013); Joss et al. (2004); Joss et al. (2006); Suárez et al. (2012); Suarez et al. (2010). The relatively high values (stars in Fig. 2a) are obtained from Suarez et al. (2010). The dashline represents a somewhat arbritrary boundary between slowly (Kbiol < 1 L/(g-TSS d)) and easily (Kbiol > 1 L/(g-TSS d)) biodegradable micropollutants. (b) Biological pathways involved in micropollutant transformation in MBRs.

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Fig. 3. Micropollutant removal efficiency as a function of (a) SRT and (b) HRT. (c) Typical SRT, HRT and F/M ratio used for MBRs treating micropollutants. In Fig. 3b the C/N ratio stands for the carbon to nitrogen ratio of the influent wastewater. Data in Figs. 3a and b are collected from Maeng et al. (2013) and Boonnorat et al. (2016) respectively. Data in Fig. 3c include the studies by Besha et al. (2017), Boonnorat et al. (2016), Delgado et al. (2011), Gurung et al. (2017), Harb et al. (2016), Kruglova et al. (2016), Sanguanpak et al. (2015) and Zolfaghari et al. (2015).

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Fig. 4. A comprehensive procedure evaluating the interaction of micropollutants with microorganisms and the influences on MBR performance. The left module represents the exposure procedure to determine the acute toxicity of the selected compound, the middle module illustrates the responding criteria (Zhang et al., 2018), and the right module depicts a standard MBR setup to investigate the chronic effects on the process performance. ROS is the acronyms of reactive oxygen species. Previous studies have described how to determine the microbial products (Han et al., 2013), how to characterize the adhesion and fluidity properties (Mei et al., 2014), how to determine the biodegradation and nitrogen metabolism kinetics (Han et al., 2016; Song et al., 2018), and how to evaluate the long-term performance in a flow system (Wang et al., 2014a).

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Table 1. Four case studies on micropollutant removal in pilot-scale/full-scale MBRs treating municipal wastewater. Configurationa

AOOA

Capacity

743 m3/d

HRT

SRT

MLSS

(h)

(d)

(g/L)

36-

25

41

Performance

Ref.

8.5 ±

Except caffeine, estrone and

Phan et al.

0.7

triclosan, micropollutant

(2015b)

concentrations in effluent < Australian Guidelines for Water Recycling: Augmentation of Drinking Water Supplies. O

800

24

equivalent

10 -

6.8 -

Log reduction of microbial

van den

15

8.1

indicators ranged between 4.9 - 5.9

Akker et

log10 units, comparable to values

al. (2014)

persons

reported in pilot-scale studies. O

3 m3/d

35

25 -

5.3 -

Removal efficiencies of bisoprolol,

Gurung et

30

9.8

diclofenac and bisphenol A = 65%,

al. (2017)

38% and > 97%, respectively; carbamazepine was not efficiently removed (℃89% to 28%). -

800 equivalent persons

24

10 -

7.5 -

Higher percentage removal via

Trinh et al.

15

8.5

biotransformation and significantly

(2016)

lower percentage removal via adsorption to biomass for 17βestradiol, estrone, bisphenol A and

50

triclosan during summer (high temperature). a. AOOA: Anoxic, aerobic, aerobic, anoxic; O: Aerobic

51

Highlights ►The increasing occurrence of micropollutants poses threats to water reclamation. ►Removal in MBRs relate to sorption, biodegradation and membrane separation. ►Sorption efficiency depends on the hydrophobicity and ionogenicity of compounds. ►SRT, HRT, ORP, etc. influence the biological degradation performance. ►Efficient membrane separation can be achieved by technology integration.

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