# Removal of organic micropollutants in anaerobic membrane bioreactors in wastewater treatment: critical review

Michael Lim, Dominique Steyer Patureau, Marc Héran, Geoffroy Lesage, Jeonghwan Kim

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Michael Lim, Dominique Steyer Patureau, Marc Héran, Geoffroy Lesage, Jeonghwan Kim. Removal of organic micropollutants in anaerobic membrane bioreactors in wastewater treatment: critical review. Environmental Science: Water Research and Technology, 2020, 6 (5), pp.1230-1243. 10.1039/c9ew01058k. hal-02650858

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# Removals of Organic Micropollutants in Anaerobic Membrane Bioreactor in Wastewater Treatment: Critical Review

Michael Lim¹,², Dominique Patureau³, Marc Heran⁴, Geoffroy Lesage⁵ and
Jeonghwan Kim¹*

¹Department of Environmental Engineering, Inha University, Incheon, Republic of Korea

²Department of Chemical and Materials Engineering, The University of Auckland, Auckland, New Zealand

³LBE, Univ Montpellier, INRA, Narbonne, France

⁴IEM, Univ Montpellier, CNRS, ENSCM, Montpellier, France

*corresponding author: jeonghwankim@inha.ac.kr

## Abstract

Anaerobic membrane bioreactor (AnMBR) is a promising technology for achieving an energy-saving or even energy-positive wastewater treatment process as it produces high effluent quality and renewable energy in the form of methane. Nevertheless, concerns on inflowing organic micropollutants (OMPs) caused by various human and industrial activities into AnMBR system are proliferating. The extent to which the removal of OMP in the AnMBR is understood should vary greatly depending upon the removal pathways mainly involved such as sorption into biomass, transformation, or membrane filtration. This review paper describes the fate and removal mechanisms of OMPs in AnMBR system. Although the overall performance of AnMBR treating various wastewaters has been observed under such reduced conditions, understanding and modeling the removal mechanisms of OMPs in this type of reactor still requires many works. Elucidating the removal mechanisms of OMPs will lead to the improvements in designs and operations of AnMBR system while optimizing performances and saving energy.
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**Keywords**

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Organic micropollutants, anaerobic membrane bioreactor, wastewater, biodegradation, adsorption, absorption, membrane separation

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**Introduction**

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Organic micropollutants (OMPs) in domestic wastewater are generated mainly by various human and industrial activities, covering a wide range of contaminants including pharmaceuticals, pesticides, personal care products (PCP), industrial chemicals, hormones, and other emerging compounds. While the OMPs are present at trace amounts in the wastewater ranging from ng/L to µg/L, tremendous efforts have been made to better understand the fate and transport of the OMPs in water bodies significantly with considering their significant risks to human health and eco-environmental security¹. Those risks include short-term or long-term toxicity, environmental persistence, antibiotic resistance, endocrine disruption etc². Nevertheless, the individual or collective behavior of the OMPs in domestic wastewater treatment plant (WWTP) require much attention since the information on their removal behaviors through the biological treatment process are still limited particularly under anaerobic conditions.

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Interests in anaerobic membrane bioreactor (AnMBR) are growing in domestic wastewater treatment or high strength wastewater rapidly since it enables synergistic effects on wastewater managements and resource recovery³, ⁴. The AnMBR is to combine anaerobic bioreactor with membrane filtration. The key advantages offered by the AnMBR process are to uncouple hydraulic retention time (HRT) and solid retention time (SRT). Compared to conventional aerobic biological treatment such as activated sludge process, the AnMBR

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allows (1) low sludge production (up to 90 %)⁵, (2) low energy consumption and (3) production of bioenergy in the form of methane⁶ as well as (4) production of high effluent quality (in terms of suspended solids, turbidity and microorganisms) thanks to membrane filtration. Nevertheless, understanding the removal of OMPs in AnMBR system still requires many works.

Both sorption and biotransformation play primary roles in the removals of OMPs during operation of AnMBR⁷. Sorption of OMPs from the aqueous phase to biomass should occur due to hydrophobic interactions and electrostatic interactions (cation bridging and exchange) ¹, ⁸. Biotransformation is the process whereby microbes decompose organic pollutants. In AnMBR, porous membrane such as microfiltration (MF) or ultrafiltration (UF) is often used to retain the biomass from wastewater. However, rejection efficiency of the OMP through MF or UF may not be very high because the size of OMP molecule is often smaller than the pore size of membrane which is in the range from 0.1 to 0.01 µm. Nevertheless, it allows the retention of colloidal organic carbon which are largely bound to OMP. During membrane filtration, fouling caused by the deposit of organic matter on membrane and/or within membrane pores is an unavoidable phenomenon. It is also known that higher concentration of OMPs in wastewater can result in the formation of denser and more compact structure of fouling layer on membrane⁹. This is because that the fouling layer formed on membrane surface can often play a role as a secondary membrane to improve OMPs rejection¹⁰.

Nowadays, there has been an upsurge of interests in AnMBR in which biomass carriers are suspended for domestic wastewater treatment¹¹. The carriers are added into the AnMBR not only for providing extra surface area for the growth of fixed biomass but also for
72 inducing mechanical membrane cleaning thanks to scouring actions. In anaerobic fluidized
73 bed membrane bioreactor termed as AFMBR, granular activated carbon (GAC) or powdered
74 activated carbon (PAC) or polyurethane sponge are often used as fluidized media⁴, ¹¹,
75 ¹² These media are fluidized by recirculating the bulk wastewater alone through the reactor
76 without biogas sparging. As a result, operational energy can be reduced significantly⁴. The
77 AFMBR is reported to have superior performance in the removal of pharmaceuticals
78 compared to that from the conventional activated sludge process treating a real domestic
79 sewage¹³. However, understanding the pathways on degrading the OMPs through AFMBR
80 system needs more experimental works and practical applications with real wastewaters¹⁴.

81 In this paper, the removal mechanisms of OMPs in the biological wastewater
82 treatment process are critically reviewed, mainly focusing on anaerobic systems. This review
83 is expected to provide a vehicle by which an understanding of OMPs fate in biological
84 wastewater treatment can be used to develop a new design of the AnMBR reactor to improve
85 the OMPs removal performance.

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Organic micropollutants in AnMBR

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*Organic micropollutants (OMPs) issues in wastewater*

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Many chemicals from human activities are present in trace amounts in natural
89 environments (e.g., plasticizers, detergents, hydrocarbons, solvents, pesticides, cosmetics,
90 and drugs). Their concentrations rarely exceed a few hundred nanograms per liter. While
91 the risks associated with chronic exposure to these substances are still widely discussed by
92 scientists, there is considerable research showing that at these concentrations, some
93 substances will affect the behavior of aquatic organisms and human health. As a result, these
94 substances are called "micropollutants" Wastewater treatment plants (WWTPs) are
considered to be one of the primary sources of micropollutants to aquatic environments.

Since the year of 2000, the European regulation, for example, has encouraged the reduction of emissions in application of the objectives set by the Water Framework Directive (WFD).

At the European level, a list of so-called "priority" substances (whose discharges are to be reduced) and "dangerous priority" substances (whose discharges are to be deleted) was published in 2001 and updated in 2008 and 2013. In parallel, many research teams are studying other micropollutants, so-called "emerging" substances (not yet regulated), due to a lack of knowledge about exposure levels and/or their toxicity to aquatic environments.

Considerable diversity of organic micropollutants is measured in urban wastewaters, being the mirror of the vital production and consumption of chemicals of human beings. These OMPs include a large variety of substances families (volatile organic compounds, detergents, plasticizers, flame retardants, pesticides, hormones, solvents, pharmaceuticals, PCPs, Polycyclic aromatic hydrocarbon (PAH), surfactants etc) with various physico-chemical properties (hydrophilic/ hydrophobic, from low to high molecular weight) and their own removal pathway (Table 1)¹, ⁴, ¹¹, ¹⁵⁻¹⁷.

The WWTPs are designed for the removal of organic matter and nutrients. Concomitantly they are able to remove some of these organic micropollutants thanks to (bio) transformation and sorption onto sludge¹⁸⁻²¹. However, some of them are recalcitrant, and effluent discharges from the WWTP constitute one of the main inputs of OMPs into the environment¹, ¹⁷, ²². In order to reduce these discharges, tertiary treatments like advanced oxidation processes (AOPs), the addition of activated carbon, membrane filtration processes and biological treatments such as wetlands, biofilters, algae reactor, MBR, MBBR, etc., were designed and operated from lab-scale to full-scale¹⁵, ¹⁹, ²³⁻²⁹. It seems that the main removal
mechanisms along WWTP are transformation (biotic or abiotic) and sorption (onto sludge, carriers, membranes) and that operational parameters such as redox conditions, HRT, SRT, temperature, pH, type and quantity of (co)substrates play a significant role onto biotransformation by driving (a) the diversity of the microbial community and the metabolic pathways (long SRT, low to high redox may increase microbial and metabolic diversity), and (b) the microbial activity (higher temperature may increase the biological kinetics). Concentrations of OMPs in the influent and effluent of WWTP vary spatially and temporally (regional, dayly, seasonal) according to industrial and domestic uses¹. The spectrum of molecules found in sewage is very wide as exemplified in Table 1³⁰. This includes pharmaceuticals, endocrine-disrupting chemicals (EDC), personal care products, surfactants, pesticides, etc, with a large spectrum of physico-chemical properties. Some of them can provide negative impacts on the wild-life in water bodies such as feminization of fish and mussels³¹. The concentrations found in influents range from the order of ng/L for pharmaceuticals¹⁷ to mg/L particulary for detergents³². In conventional wastewater treatment plants, the concentrations are generally lower in secondary effluent than those in the primary effluent ³³. For some pharmaceuticals, however, the concentrations in secondary effluents can be higher than those detected in the primary effluent due to the deconjugaison or the production of by-products through aerobic biological transformation³⁴. which are generally found in domestic sewage.

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<table><caption>**Table 1.** Types of organic micropollutants ¹, ³⁵</caption><thead><tr><td>OMPs</td><td>Classes</td><td>Major Sources</td><td>Examples</td></tr></thead><tbody><tr><td>Pharmaceuticals</td><td>Antibiotic, antidiabetic, analgesic, anticonvulsant, stimulant, veterinary drug</td><td>Excretion, hospital effluents, farmland waste</td><td>Acetaminophen, androstenedione, amoxicillin, cbz, dcf, ibp, keto, penicillin, smx, paracetamol, timolol, salicylic acid, etc.</td></tr><tr><td>Personal care products (PCPs)</td><td>Antiseptic, disinfectant, fragrance and synthetic musk, stimulant, UV filter, insects repellent</td><td>Shower, swimming, bathing, shaving, or industrial waste</td><td>Benzophenone, caf, diltiazem, chlorophene, triclosan, methylbenzylidene, chlorophene, tonalide, etc.</td></tr><tr><td>Steroids and hormones</td><td>Endocrine disruptive chemicals (EDCs)</td><td>Excretion, hospital effluents, farmlands, aquaculture</td><td>Estradiol, estrone, diethylstilbestrol, progesterone, testosterone, etc.</td></tr><tr><td>Pesticides</td><td>insecticide, herbicide, fungicide</td><td>Domestic and agriculture uses</td><td>Diuron, mecoprop, MCPA, terbuthylazine, etc</td></tr><tr><td>Detergent, surfactant</td><td>Cationic, anionic, non ionic, perfluorinated compounds (PFCs)</td><td>Bathing, laundry, dishwashing, households, dilutants, dispersants</td><td>Alkylphenol ethoxylates, alkylphenols (nonylphenol and octylphenol), perfluorooctane sulfonate, perfluorooctanoic acid</td></tr><tr><td>Others</td><td>Plasticizer, flame retardant, hydrocarbon, solvant</td><td>Leaching out of the material, improper cleaning, run-off from the garden, roadway, etc</td><td>Phthalates, polybrominated compounds, dioxin and furans, polycyclic hydrocarbons, trichloroethylene, benzene, toluene, etc</td></tr></tbody></table>
Removal pathways of organic micropollutants

According to micropollutant characteristics such as hydrophobicity, molecular weight, volatile properties, biodegradation, and aromaticity, etc., several removal pathways allowing their removal from the liquid phase can be considered as shown in Table 2. The figure 1 addresses the different items listed in Tab.2 in the global removal pathways in anaerobic membrane bioreactor, which is discussed in more detail below.

<table><caption>Table 2. Removal pathway of organic micropollutants (OMP) ¹, ³⁵</caption><thead><tr><th></th><th>n°1</th><th>n°2</th><th>n°3</th><th>n°4</th><th>n°5</th></tr></thead><tbody><tr><th>Pathway</th><td></td><td></td><td></td><td></td><td></td></tr><tr><th>Mechanisms</th><td>Coagulation/Flocculation.<br>Sorption<br>Precipitation</td><td>Volatilization</td><td>Biological<br>oxidation<br>(Biodegradation)</td><td>Chemical<br>oxidation</td><td>Liquid/Liquid<br>separation<br>(Membrane)</td></tr><tr><th>Variables</th><td>Partition coefficient<br>Solubility constant :<br>K<sub>d,OMP</sub> (L<sup>3</sup>.M<sup>-1</sup>), S</td><td>Henry's law<br>constant : k<sub>H</sub><br>(ML<sup>-4</sup>T<sup>-2</sup>N<sup>-1</sup>)</td><td>Biodegradation<br>Kinetics<br>constant :<br>k<sub>bio,OMP</sub> (T<sup>-1</sup>)</td><td>Kinetics<br>constant :<br>k<sub>03,OMP</sub> (M<sup>-1</sup>.s<sup>-1</sup>)</td><td>Molecular<br>weight<br>Charge :<br>MW<sub>OMP</sub></td></tr><tr><th>Influencing<br>parameters</th><td>Particle size and charge,<br>concentration and type of<br>organic matter,<br>hydrophobicity of<br>compounds</td><td>Gas sparging</td><td>Composition of<br>microbial<br>community,<br>redox conditions<br>SRT, HRT</td><td>Oxidant level</td><td>Permeate<br>flux,<br>membrane<br>fouling<br>SRT, HRT</td></tr><tr><th>Strong<br>influence of<br>the pathway<br>on</th><td>PAH, PCB, PBDE,<br>fluoroquinolones</td><td>Naphthalene,<br>trichloroethylene</td><td>Paracetamol,<br>hormones, ibp</td><td>Cbz, dcf</td><td>smx, cbz,<br>trim,<br>amitriptyline</td></tr></tbody></table>

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<figure><img src="image_8.png" /><figcaption>**Fig 1.** Removal pathway of organic micropollutant in anaerobic membrane bioreactor</figcaption></figure>

*Sorption to biomass or adsorbents*

Sorption to the biomass has been suggested as one of the primary mechanisms affecting the removal of the OMPs in anaerobic treatment³⁶. In fact, the sorption of OMP to biomass should be differentiated into two mechanisms : (1) Sorption in which the OMPs move from aqueous phase and (2) penetration into the lipophilic cell membrane, which is driven by their hydrophobicity. The OMPs are also sorbed onto surface of biomass due to electrostatic interactions between the biomass with negative surface charge and OMPs with positive surface charge³⁷ or negative one thanks to cation bridging. Suarez *et al.* (2008) reported that the sorption of OMPs by the biomass could be determined by the value of the Octanol-Water partitioning coefficient (K<sub>ow</sub>). In case of ionizable compounds, the K<sub>ow</sub> can be modified into distribution partitioning coefficient (D<sub>ow</sub>) by taken in account the acid
176 disassociation constant (pKa) as shown in Eq.(1)³⁸. On the other hand, field trials make it
177 possible to obtain the solid-water partitioning coefficient (K<sub>d</sub>), also known as the real ratio
178 between concentration in solid and liquid phases at equilibrium condition. This coefficient is
179 used to predict the fraction of the OMPs sorbed on the biomass³⁸, as shown in Eq.(2).

$$
180 \quad D_{ow} = \frac{K_{ow}}{1+10^{pH-pKa}} \quad \text{or} \quad D = \frac{K_d}{1+10^{pH-pKa}} \quad (1)
$$

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$$
183 \quad K_d = \frac{C_s}{C_L \times T_{SS}} \qquad (2)
$$

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185 Where:

186 $K_{ow}$ : octanol-water partitioning coefficient

187 $K_d$ : solid-liquid partition coefficient (L/g)

188 $X_{TSS}$ : Concentration of volatile suspended solids (g/L)

189 $C_s$ : the concentration in the solid phase (µg/L)

190 $C_L$ : the concentration in the liquid phase (µg/L)

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192 In 2010, Barret *et al.* described a three-compartment model for the sorption of OMPs into
193 sludge. They distinguished OMPs present in the sludge into; the freely dissolved (C<sub>free</sub>,mg/L),
194 the sorbed to Dissolved and Colloidal Matter (C<sub>DCM</sub>, mg/g<sub>DCM</sub>), and the sorbed to particles
195 (C<sub>part</sub>, mg/g<sub>PART</sub>) <sup>39</sup>. This three compartment model is more adapted to a membrane separation
196 step (AnMBR). Indeed, the retention of DCM depends on molecular cut-off of membrane and
197 properties which provides precious information to predict OMP fate in AnMBR more
198 accurately. At equilibrium condition, the three-compartment system can be explained by
199 equations below:
$$K_{part} = \frac{C_{part}}{C_{free}} \qquad (4)$$

$$K_{DCM} = \frac{C_{DCM}}{C_{free}} \qquad (5)$$

Then, the concentration of apparently dissolved micropollutant (C_L), which is the sum of freely dissolved and sorbed to DCM (C_L = C_free + C_DCM [DCM]).

$$K_d = \frac{C_{part}}{C_L} = \frac{C_{free} \cdot K_{part}}{C_{free} + C_{DCM} [DCM]} = \frac{K_{part}}{1 + K_{DCM} [DCM]}$$

Where:

[DCM] : Dissolved and Colloidal Matter (g_DCM/L)

K_DCM : Equilibrium constant of OMPs sorption to DCM (L/g_DCM)

K_part : Equilibrium constant of OMPs sorption to particles (L/g_part)

Thus, when the separation step is based on an ideal membrane (Total Retention of [DCM]), the concentration of OMP in the effluent is equal to C_free (C_free = C_L - C_DCM [DCM]) whereas the value of C_L is reached when a settler is used.

It has been found that the OMP removal by sorption not only depends on their physicochemical characteristics but also on sludge's characteristics, such as the biomass conformation and particle size, as well as operational conditions³⁹. Consequently, the AnMBR enhanced sorption and biotransformation of some OMPs. In fact, the physicochemical characteristics of the membrane will enhance the sorption capacity by adding extra solid-liquid interface. The hydrophobic membranes are indeed able to retain lipophilic OMPs, whereas charged membranes can retain OMPs due to electrostatic interaction¹⁸.
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Carrier materials are also often added into an AnMBR to facilitate biofilm growth and reduce membrane fouling. Both of them can be achieved by providing high specific surface area and mechanical scouring actions, which is driven by fluidizing the media along the membrane surface. Recently, activated carbon particles such as PAC or GAC have been mainly investigated as carrier materials or fluidized media in AnMBR. The addition of PAC or GAC into the AnMBR thus helps increase removal efficiency of soluble micropollutant. It was found that about 100 mg/L of PAC dosage played a crucial role in adsorbing organic contaminant to overcome competitive adsorption with other organics present in bulk wastewater<sup>40</sup>. Xiao et al. (2017) investigated the removal efficiency of the pharmaceutical compounds in AnMBR with and without the addition of PAC<sup>41</sup>. The PAC addition improved the removal of the pharmaceutical compounds such as Sulfamethoxazole (Smx) and Triclosan (Tcs) during operational period. This enhancement was explained by the fact that a local substrate concentration could be increased in the mesoporous structure not only provided by carbon particles but also to their high adsorption capability. Therefore, the biodegradation became more thermodynamically favorable<sup>41</sup>. However, it was also found that the removal efficiency of Carbamazepine (Cbz) and diclofenac (Dcf) was very low, which is less than 5 % because those chemicals are refractory against biodegradation<sup>41</sup>. Similar results were found by Wei et al., (2016) demonstrating the limited adsorption capacity provided by the GAC for the removal of OMPs during extended operational time with AnMBR<sup>40</sup>. Terzyk et al., (2003)<sup>42</sup> claimed that the abiotic reaction between phenol and PAC should affect irreversible adsorption<sup>40, 42</sup>. The abiotic transformation itself can be occurred by several steps. Firstly is the creation of strong complexes between the surface lactone and carbonyl groups of the carbon and phenolic group. Secondly is polymerization by the *superox* ion formed during the oxygen adsorption into PAC. However, oxygen is very limited in the AnMBR, thus
strong complexes between Cbz and Dcf and the surface groups of the carbon particles may be dominant in transformations of those micropollutants. During AnMBR operation, the carbon particles should also be occupied by the OMPs, thus mitigating the biotransformation of Cbz and Dcf <sup>42</sup>. On the other hand, the removal of the OMPs in the AnMBR can occur initially after adding the PAC followed by a gradual increase of OMP concentration to a certain level. As operation time progressed, the surface groups of PAC were occupied by the OMPs, thus decreasing biotransformation of Cbz and Dcf <sup>42</sup>.

On the other hand, Wei *et al.* (2016) <sup>40</sup> reported that the reduction of OMPs only occurred initially after the PAC addition in AnMBR followed by a gradual increase of OMPs concentration to the level before PAC addition. It was also predicted that the low dosage of PAC (100 mg/L) played a significant role for this limited adsorption capability considering the sorption competition with other bulk organics<sup>40</sup>. Granular activated carbon (GAC) particles were added as carrier materials for the growth of biofilm and physical scouring to clean membrane surface. More than 90 % in removal efficiency was achieved at 5 hr of total HRT <sup>43</sup>. Sorption into GAC particles as much as biofilm formation on GAC and its associated biodegradation form the basis of the two crucial pathways responsible of AFMBR pharmaceuticals removal<sup>34</sup>.

## Volatilization

Volatilization of OMPs from dissolved to gas compartment is mainly determined by physicochemical properties of the OMPs. The volatility of OMPs can be characterized by the value of Henry's law constant (k<sub>H</sub>) in which k<sub>H</sub> value ranging from 10<sup>-2</sup> to 10<sup>-3</sup> mol/(m<sup>3</sup>.Pa) corresponds to the high tendency of volatilization<sup>1</sup>. Suarez *et al.* (2008)<sup>38</sup> reported that volatilization of OMPs from estrogens and pharmaceuticals group is nearly negligible<sup>38</sup>.
However, it can also act as the major removal pathway for semi-volatile or non-biodegradable OMPs. In addition, volatilization behavior is intensified in the wastewater treatment with the biogas production or membrane sparging and temperature. In fact, Feigenbrugel et al. (2004)⁸⁴ reported that the behaviour of Henry's law constants as a function of temperature can be expressed with a Van't Hoff equation. Thus, when the temperature increase from 25°C to 45°C, the volatilization also will increase by a factor of 8.44 for metalachlor and 12 for diazinon. The equation of an equilibrium condition in the interface between the atmospheric and the wastewater surface is described as below in Eq (6)³⁸, ⁴⁴.

$$R_{\text{vol.surf}} = -V K_L a_{\text{surf}} C_f \frac{H_c}{H_c + \left(\frac{k_L}{k_G}\right)_{\text{surf}}} \quad (6)$$

In which:

R<sub>vol.surf</sub> : Rate of Volatilization (gd<sup>-1</sup>)

V : Volume (m<sup>3</sup>)

H<sub>c</sub> : dimensionless Henry's law constant of compound (expressed as m<sup>3</sup> of wastewater/m<sup>3</sup> of air)

C<sub>f</sub> : concentration of compound (mg L<sup>-1</sup>)

a<sub>surf</sub> : specific interface surface area (m<sup>-1</sup>)

k<sub>L</sub> : liquid transfer mass coefficient (m d<sup>-1</sup>)

k<sub>G</sub> : gas transfer mass coefficient (m d<sup>-1</sup>)

K<sub>L</sub>a<sub>surf</sub> : the overall surface-desorption gas-transfer coefficient (d<sup>-1</sup>)

(k<sub>L</sub>/k<sub>G</sub>)<sub>surf</sub> : Ratio of the liquid-phase mass-transfer coefficient to the gas-phase mass-transfer coefficient for the surface volatilization

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## Biological transformation

291 Biotransformation of OMPs has been explored because microbial cells should have the ability to breakdown the molecular structure of OMPs either totally (which is called
mineralization) or partially with the production of transformation products (TPs). These
biotransformations can be divided into two reactions: (1) the metabolic reactions where the
OMPs are used as carbon (anabolism) and/or energy (catabolism) sources for their cell
development (growth, maintenance)²¹ and (2) the co-metabolic reactions where the OMPs
could not sustain growth but are transformed by side reactions catalyzed by broad-spectrum
non-specific enzymes, in this case a primary substrate is needed⁴⁵. In the case of organic
micropollutants present at very low concentration in WWTP, it is expected that the OMPs
could not sustain growth and that the primary substrate is needed. Models were developed in
order to describe this co-metabolic reaction based on the original model developed by Criddle
(1993)⁴⁶ and underlying the beneficial effect of the primary substrate⁴⁷⁻⁴⁹. Despite the various
redox conditions occurring in WWTP, the biotransformation under aerobic conditions has
been more studied than the one under anoxic or strictly anaerobic conditions⁵⁰, ⁵¹. Indeed, the
microbial and functional diversity of anaerobic ecosystems met in WW, sludge and livestock
treatments, may lead to the biotransformation of OMPs³⁶, ³⁹. The most observed reactions are
the reductive dehalogenation and the cleavage of ether bonds mainly observed with simple
molecules using pure or enrichment cultures³⁹, ⁵⁰, ⁵². However in complex anaerobic systems
treating sludge or livestock effluent (lab/pilot/industrial-scale), many compounds have been
shown to be well removed under these reduced conditions like naproxen, atenolol, loratidine,
miconazole, tramadol, domperidone, azithromycine, trimethoprim, tylosine, some
sulfonamides (table 3), however (i) large variability in the removal rates depending on
sampling and experimental conditions, (ii) no apparent relationship with their sorption
capacity (Eq. 1), LogK<sub>d</sub> or LogD explaining the partition between aqueous and particulate
phases but not transformation which would be barely linked to their molecular structure and
the presence of electron donating functional groups like -NH<sub>2</sub>, -OH, -CH<sub>3</sub> and (iii) no
knowledge on their real elimination pathways (abiotic/biotic, major or minor structural modification) since transformation products are rarely identified. Very recently, Spielmeyer et al., (2017)⁵³ have quantified transformative products (TPs) and assessed their antimicrobial activity for three sulfonamides like sulfadimethoxine, sulfamethoxypyridazine, sulfamethoxazole (SMX) during anaerobic digestion of manure under batch reactor spiked with the molecules [43]. Two TPs were isolated by coming from demethylation of sulfadimethoxine and were also subsequently removed. Comparison with autoclaved systems supported the hypothesis of biotic reactions. For sulfamethoxypyridazine, one TP was identified also coming from demethylation that was not further degraded during the batch reactor operation. SMX seemed to be biotransformed through the hydrogenation of the double bond and reductive cleavage of the heterocycle; a hydrogenation step was also mentioned by Alvarino et al., (2016)¹⁸ for SMX. Another TP was also detected but coming from abiotic transformation. They also observed transformation of CTC to iso-CTC, transformation that occurred abiotically. Mono and polycyclic aromatic compounds (benzene, phenol, low molecular weight PAH) anaerobic transformation were also quite well documented in literature with description of metabolic routes and isolation of degrading microorganisms⁵⁴.

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<table><caption>**Table 3.** The range of OMPs removal under anaerobic conditions. Underlined OMPs are the one found in only one class of biological removal ( 28, 31, 39, 53, 55-73)</caption><thead><tr><th>Class of removal</th><th>&lt;30</th><th>&lt;30-70&gt;</th><th>&gt;70</th></tr></thead><tbody><tr><td>Sludge<sup>1</sup></td><td>para, cbz, dcf, ibp, flx, acid salicylic, gem, ofl, nor, cip, LAS, NP, NP2EO, PAH, PCB, E1, E3, T, αEE2, αE2, βE2, DEHP, BBP, DEP, BPA, ahtn, hhcb, triclosan, triclocarban, diuron, benzotriazole, clozapine, benzophenone, iopromide, bisoprolol</td><td>pfoa, pfos, para, cbz, propra, smx, cefo, esci, lido, vera, citalopram, keto, ibp, dcf, diazepam, roxi, ctc, ofl, nor, cip, LAS, NP2EO, PCB, E1, E3, T, αEE2, αE2, βE2, DEHP, BBP, DEP, DnBP, BPA, ahtn, hhcb, triclosan, triclocarban</td><td>pfos, para, cbz, propra, smx, azi, cefo, esci, lido, lora, mico, trama, vera, domp, dcf, ibp, ate, caf, trim, nap, oxybenzone, roxi, otc, flx, citalopram, furosemide, clofibric acid, keto, nor, cip, NP2EO, NP1EO, E2, E1, αEE2, DEP, DnBP, BPA, ahtn, hhcb, triclosan</td></tr><tr><td>Livestock effluent<sup>2</sup></td><td>sulfathiazole, sulfamethazine, sulfadiazine, sulfaguanidine, sulfamerazine, sulfapyridine, monensine, doxycycline, tetracycline</td><td>oxytetracycline, sulfachloropyridazine, sulfathiazole</td><td>smx, sulfamerazine, sulfadiazine, sulfadimethoxine, sulfamethoxypyridazine, trimethoprim, tylosine, florfenicol, ampicillin, chlortetracycline</td></tr></tbody></table>

**Note:** Cip: Ciprofloxacin, nor: Norfloxacin, ofl: Ofloxacin, para: Acetaminophen/paracetamol, az i: Azithromycin, cbz: Carbamazepine, cefo: Cefoperazone, domp: Domperidone, esci: Escitalopram, glyben: Glybencyclamide, iver: Ivermectine, lido: Lidocaine, lora: Loratadine, mico: Miconazole, propra: Propranolol, smx: Sulfamethoxazole, trama: Tramadol, vera: Verapamil, pfoa: perfluorooctanoate, pfos: Perfluorooctane sulfonate, E1: Estrone, E3: Estriol, T: Testosterone, αE2: 17-α-estradiol, βE2: 17-β-estradiol, αEE2: 17-α-ethinylestradiol, NP: nonylphenol, NP1EO: Nonylphenol monoethoxylate, NP2EO: Nonylphenol diethoxylate, OP: octylphenol, BBP: Benzyl butyl phthalate, DEHP: Diethyl hexyl phthalate, DEP: diethyl phthalate, DnBP: Di(n)butyl phthalate, LAS: linear alkylbenzene sulfonate, dcf: diclofenac, ibp: ibuprofen, keto: ketoprofen, Roxi: R oxithromycin, Ate: Atenolol, Caf: Caffeine, Trim: Trimethoprim, Nap: Naproxen, otc: oxytetracycline, ctc: chlortetracycline, flx: fluoxetine, gem: gemfibrozil, PAH: Polycyclic aromatic hydrocarbons, PCB: polychlorobiphenyls, ahtn, hhcb: Tonalide, galaxolide, BPA: Bisphenol A, 4-OP: 4-(1,1,3,3-tetramethylbutyl)phenol, t-NP: technical-nonylphenol, 4-NP: 4-n-nonylphenol
It is assumed that biotransformation occurs only in the sludge phase when the OMP
is transferred from the liquid phase through several mechanisms. Firstly, sorption-desorption
can limit the biotransformation of a highly hydrophobic compound, and secondly small size
of OMPs allows them to diffuse through outer membrane in the microbial cell.
Biotransformation rate is strongly dependent upon the OMPs physicochemical and structural
properties, the enzymatic activities, the microbial diversity, and the environmental and
operational parameters ³⁶. The anaerobic digestion of sludge can be described by using a two-
steps model; first hydrolysis to particulate matter (Xs) followed by biodegradation of soluble
substrate (Ss) to biogas ⁴⁷.

$$
S_p \xrightarrow{k_{hyd}} S_s \qquad (7)
$$

$$
S_s \xrightarrow{\mu} X + CH_4 + CO_2 \qquad (8)
$$

In which, biomass growth rate is linked to soluble substrate uptake and modeled with Monod-
type kinetics:

$$
\mu = \mu_{\max} \frac{S_g}{K_s + S_g} \qquad (9)
$$

Where μmax (1/d) is the maximum bacterial growth rate, and Ks (gCOD/L) is the half-saturation constant associated with the soluble substrate Ss⁴⁷.

In 1993, Criddle modeled the cometabolism equation as shown in Eq. (1) between a growing
and non-growing substrate by biomass⁴⁶. This equation was then modified by Delgadillo-
Mirquez *et al.* (2011) in which OMPs are considered as the non-growing substrate ⁴⁷.

$$
r_{\text{bio}} = \left( T_c^{\frac{\mu}{Y}} + k_c \right) \left( \frac{c_{\text{bio}}}{k_{sc} + c_{\text{bio}}} \right) X \quad (10)
$$

Where
378 $r_{bio}$ = biotransformation rate constant

379 $T_c$ = OMPs transformation capacity ($\mu g_{OMP}/g_{COD-Ss}$) standing for cometabolic interaction
380 between the soluble substrate and OMPs metabolism,

381 $k_c$ = the maximum specific rate of OMPs in the absence of primary substrate ($\mu g_{OMP}/g_{COD-x.d}$)

382 $K_{sc}$ = the half saturated constant of OMP in monod formalism ($\mu g_{OMP}/L$)

384 $\mu$ = the growth rate (1/d)

385 $Y$ = the growth yield ($g_{COD-x}/g_{COD-s}$)

386 $C_{bio}$= the concentration of soluble substrate to be biodegraded

387 Co-metabolisms is a process by which the OMPs are fortuitously biodegraded by using the
388 enzyme produced during the metabolisms of other compounds<sup>74</sup>. The co-metabolisms always
389 involve external organic compound as carbon and energy sources to partially or entirely
390 biodegrade the OMPs in wastewaters. The co-metabolisms enable the biodegradation of
391 OMPs which are far below the concentration which can be used as energy and carbon sources
392 by microbial cells<sup>74</sup>. Structure analog is a chemical compound which is similar chemical
393 structure to the target biodegradable OMP compound. Addition of structural analogue
394 compound can encourage the growth and activity of microbial cells to degrade organic target
395 compound<sup>75</sup>. Structural analogue influences biodegradation by acting as an inducer of
396 catabolic pathway and substrate for co-metabolic in biodegradation of OMPs <sup>76</sup>.

397

398 *Liquid-liquid separation*

399 Membranes in AnMBR can always provide a useful tool to reject OMPs through membrane
400 pore matrix. In most of the AnMBR system, porous membranes such as microfiltration or
ultrafiltration membrane are used directly into anaerobic bioreactor contributing to the removal of OMP as a result of size exclusion and charge repulsion where the physicochemical characteristics of the membrane and OMPs determine rejection capacity.

However, the rejection efficiency of OMPs may not be too high due to smaller size of OMPs than the pore size of membrane applied<sup>77</sup>. Alvarino et al.(2018) report that hydrophobic membrane is able to retain lipophilic OMPs, whereas charged membranes can retain OMPs by electrostatic interaction<sup>18</sup>. In fact, with pore sizes in the range 0.01-0.1µm, the membrane sieving effect concerns only pollutants with a molecular weight higher than the molecular weight cut-off (MWCO) of membrane. Nevertheless, it has been reported that the removal of micropollutants may also be achieved by MF/UF due to the rejection/sorption by the foulants that act as a secondary barrier on membranes<sup>78, 79</sup>. Moreover, membrane allows also the retention of a significant part of the DOM and its associated OMP sorbed. As a result, the overall removal efficiency via biodegradation and direct rejection by membrane were increased<sup>40</sup>. Monsalvo et al. (2014)<sup>80</sup> characterized the distribution of trace organics within the fouling layer which is formed on the membrane in AnMBR and found that 17 α-ethinylestradiol, estrone, octyphenol, and bisphenol A, were retained mostly due to the layer of membrane fouling<sup>80</sup>. Rejections of the OMPs by membrane also increase the concentration polarization on membrane surface. Therefore, threshold value of wall-concentration of OMPs can be achieved to activate their biological degradation and thus shortening their adaptation time. Thus, the presence of a fouling layer on the membrane surface in AnMBR affect the rejection of OMPs due to the pore blockage, cake enhanced concentration-polarization, the change in surface charge of membrane and hydrophilicity<sup>78</sup>. Membrane fouling in AnMBR can be classified into (1) biofouling, (2) organic fouling, (3) colloidal fouling and (4) inorganic fouling<sup>81</sup>. Each type of membrane fouling formed can result in the different
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rejection efficiency of OMPs. The presence of organic fouling on membrane surface
improved the rejection of OMPs due to the enhancement of steric hindrance by the organic
cake layer on membrane⁸². However, sometimes, the rejection can be decreased after organic
fouling or colloidal fouling is formed on membrane⁷⁹, ⁸³.

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Global removal efficiency of OMPs by AnMBR and enhanced AnMBR

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Studies on the OMP removal in AnMBR have particular relevance to biodegradation, adsorption, and membrane filtration. Dutta et al., (2014)⁴³ observed the removal of the pharmaceuticals in anaerobic fluidized bed membrane bioreactor (AFMBR) as a second stage for the treatment of the effluent produced by anaerobic fluidized bed bioreactor (AFBR) treating a municipal wastewater⁴³. Monsalvo et al. (2014)⁸⁰ investigated removal efficiency of 38 OMP compounds including pharmaceuticals, personal care products, endocrine disruptors, and pesticides in AnMBR system⁸⁰. Only 9 out of 38 OMPs were removed at higher than 90 %, and the rest of them was removed only less than 50 % efficiency. With synthetic municipal wastewater, six OMPs such as *amitriptyline*, *diphenhydramine*, *flx*, *smx*, *TDCPP*, and *trim* were biodegraded more easily higher than 80 % of removal efficiency. However, *atrazine*, *cbz*, *DEET*, *Dilantin*, *primidone* and *TCEP* showed more refractory characteristics against biodegradation yielding less than 40 % removal efficiency. *Acetaminophen*, *atenolol* and *caf* required a prolonged time of about 45 hr for their adaptation to AnMBR environment.

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The removal efficiencies of OMPs under anaerobic conditions should be influenced
by operational parameters. The longer solid retention time (SRT) can often result in higher
biomass and microbial diversity, thus improving the biodegradation of OMPs³⁷, ⁸⁴. However,
the impact of SRT on removal efficiency was dependent upon the types of OMPs. Carballa et
al (2007) observed higher removal efficiency of pharmaceuticals by longer SRT, but no

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influence of SRT on personal care products (PCPs) during anaerobic digestion⁸⁵. The biodegradation of some pharmaceuticals was correlated with the methanogenic rate and the biomass structure which can improve intra-molecular diffusion. Reactor hydrodynamic and biomass characteristics determine the sorption efficiency. The increase in the HRT could increase the sorption of PPCPs into biomass by providing enough contact time to reach sorption equilibrium⁸⁶. Also, the OMP removal by sorption is dependent upon the biomass conformation and particle size distribution. Decreasing the HRT in AnMBR at extending SRT results in high biomass concentration, causing severe membrane fouling and deterioration of the effluent quality¹⁴. Obviously, trade-off in selecting HRT and SRT in AnMBR should exist between OMP removal efficiency and fouling mitigation. However, membrane fouling which is an inevitable phenomena in AnMBR should also be influenced significantly by operationg parameters and types of foulants present. Thus, further optimization to control strategies for AnMBR needs to be conducted from mechanistic knowledge of fouling and probably developing process models to adaptation in practice⁸⁷.

Bacterial population should be one of the key fators in the removal of OMPs. It was observed that antibiotic-type OMPs were more readily biodegraded by the anaerobic MBR rather than aerobic MBR systems⁷, ⁸⁸. This could be attributed to a lower abundance of antibiotic resistant genes in the anaerobic environment³, ⁸⁸. In addition, adsorption process into biomass is enhanced under anaerobic conditions due to higher concentration of soluble compound along with accumulation potential of micropollutants¹³, ⁸⁹. Understanding the role of anaerobic stages such as acidogenesis and methanogenesis on the removal of OMPs should be importantly considered. Carnerio et al., categorized the contribution into three groups such as OMPs with a biotransformation significantly higher in acidogenic condition, OMPs equally transformed in both stages and the OMPs with biotransformation higher in
473

acetogenic/methanogenic conditions⁹⁰. Nevertheless, more works are needed to elucidate dominant contribution of each microbial stage to the main metabolic pathways and products formed during the OMP biotransformation in AnMBR¹⁸.

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The AnMBR is often combined with other unit technology for pre or post-treatment such as ozonation⁹¹, nanofiltration⁴⁰, forward osmosis⁹², and membrane distillation ⁹³. In addition, the AnMBR is combined with carrier media such as GAC or PAC to improve the removal efficiency of the OMPs through biofilm growth⁴¹. Kaya et al., (2017)⁹¹ combined ozonation as a pretreatment for the AnMBR system treating the wastewater containing *etodolac* from pharmaceutical industry⁹¹. Although activated carbon particles provide high surface area for biofilm formation and intrinsic adsorption capability, the breakage of them during AnMBR operation play roles as potential foulants against membrane performance⁹⁴. Therefore, alternative media which are cost-effective to overcome the weakness of activated carbon will need to be developed for AnMBR. Here, the ozonation was used as a pretreatment to avoid sulfide inhibition due to oxidation of sulfate. The *Etodolac* was removed almost entirely with 99 % of removal efficiency by pre-ozonation⁹¹. Although overall reactor performance is not altered by the introduction of OMPs at low concentration, microbial community in AnMBR can be affected significantly⁸⁸. The higher removal efficiency was observed by multiple antibiotic OMPs in the AnMBR as mentioned above, but the biodegradation associated with gene expression was impacted. Ozone has been known as very strong oxidant to remove micropollutants, but intermediates by-produce can be present rather than their complete mineralization. Wang et al. observed that the AnMBR was adequate for the removal of the active ingredient of personal care products through both biotransformation and sorption phenomena onto the biosolids⁸⁰. It was also found that the AnMBR resulted in effective removal of OMPs due to the fouling layer formed on membrane

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497 which could play a role as a secondary membrane¹⁰. Removal diagram of OMP in AnMBR
498 is tabulated in Fig 2.

499

<figure><img src="image_9.png" /><figcaption>**Fig 2.** Removal diagram of OMP during anaerobic membrane bioreactor process⁹⁵</figcaption></figure>

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<caption>**Table 4. Removal efficiency (%) of OMPs utilizing hybrid membrane bioreactor process¹⁸, ¹⁹,**</caption>

31, 36, 40, 43, 53, 57-60, 64-68, 70, 73, 81, 88, 92, 93, 96, 97

<table><thead><tr><th rowspan="2">Treatment</th><th colspan="3">Removal efficiency (%)</th></tr><tr><th>Under 30 % removal<br>(Moderate Rejection)</th><th>Between 30 and 70 %<br>removal<br>(Moderate to High<br>Rejection)</th><th>Above 70 %<br>removal<br>(High Rejeciton<br>due to Biological<br>Acitivity and<br>Stripping)</th></tr></thead><tbody><tr><td>Anaerobic<br>Membrane<br>Bioreactor<br>(AnMBR)</td><td>Ate, αE2, αEE2,<br>atrazine, cbz, clozapine,<br>DEET, dcf, Dilantin, E3,<br>E1, gem, hydroxyzine,<br>ibp keto, linuron,<br>meprobamate,<br>omeprazole, primidone</td><td>βE2, Amytriptyline,<br>BPA, enalapril,<br>etiochlolanolone,<br>paracetamol,<br>triclocarban, trim</td><td>Androstenedione,<br>androsterone, caf,<br>metformin,<br>naproxen,<br>nonylphenol, OP,<br>smx, T, triclosan,<br>vera</td></tr><tr><td>Two staged<br>AFBR-AFMBR<br>+ GAC as<br>carrier material</td><td>-</td><td>-</td><td>Sulfadiazine,<br>smx,<br>sulfathiazole,<br>sulfamethazine.<br>Erythromycin-<br>H₂O,<br>clarithromycin,<br>josamycin,<br>roxithromycin,<br>tylosin, nalidixic<br>acid, flumequine,<br>pipemidic acid,<br>norfloxacin,<br>ciprofloxacin,<br>ofloxacin,<br>cephalexin,<br>cephradine, Trim,<br>cbz,<br>psychastimulants,<br>caf, ibp,<br>naproxen,<br>ketoprofen, dcf</td></tr><tr><td>AnMBR with<br>GAC/PAC</td><td>-</td><td>-</td><td>Trim, smx, cbz,<br>dcf, triclosan</td></tr><tr><td>AnMBR<br>followed by</td><td>-</td><td>-</td><td>Etodolac</td></tr></tbody></table>
<table><tr><td>ozonation</td><td></td><td></td><td></td></tr><tr><td>AnMBR followed by NF</td><td>-</td><td>-</td><td>Para, DEET, caf, atrazine, cbz, primidone, Dilantin, smx, diphenhydramine, ate, amitriptytline, TCEP, and flx</td></tr><tr><td>AnMBR followed by FO</td><td>-</td><td>-</td><td>Caf, ate, atrazine</td></tr><tr><td>AnMBR followed by MD</td><td>-</td><td>-</td><td>Caf, smx, keto, trim, para, nap, primidone, ibp, triamterene, carazolol, TCEP, dcf, cbz, gem, simazine, amitriptyline, atrazine, diuron, propylparaben, linuron, clozapine, phenylphenol A, diazinon, triclosan, triclocarban</td></tr><tr><td>SAnMBR followed by light, oxygen and microalgae treatment</td><td>-</td><td>-</td><td>4-OP, t-NP, 4-NP, BPA</td></tr></table>

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It is always attractive to combine AnMBR with high-pressure driven membrane such as reverse osmosis for wastewater reuse purpose. Table 4 summarizes removal efficiency of OMPs in various hybrid anaerobic membrane bioreactor process. Cho et al. (1999) observed that either nanofiltration (NF) or reverse osmosis (RO) membrane indicated more than 85 % of the rejection efficiency of the biodegradable or non-biodegradable OMPs present in wastewater ⁹⁸. In addition, Wei et al., (2016) reported high rejection efficiency of 80-92%
with 15 OMPs from various OMPs sources such as pharmaceuticals, personal care products, and pesticides from integrated AnMBR-NF system⁴⁰. The removal efficiency of the OMPs by integrated AnMBR-NF process (87 %) was higher than that by the NF membrane system alone. The fouling formed on the NF membrane by pretreated anionic ion exchange effluent can vary the effectiveness of the rejection of the OMPs due to their surface charge. The rejection of the OMPs, which is more positively charged, can be lower than that of the OMPs, which are more negatively charged due to electrostatic interactions between the foulants and membrane. Comparison of fouled membrane and non-fouled membrane on the rejection of OMPs has also been investigated. In spite of high removal efficiency of OMPs with RO or NF membrane followed by AnMBR, operational energy is still very high to provide very high hydraulic pressure to allow water passage through membrane. Recently, low-pressure driven membrane process with high rejection efficiency of OMPs from AnMBR effluent has been considered significantly. For instance, forward osmosis (FO) which is osmotically-driven membrane process are integrated to remove OMPs from wastewaters. Linares et al. compared the rejection of the 13 OMPs selected as target contaminants through the clean and fouled FO membrane treating secondary effluent⁹⁹. The rejection of the OMPs was increased except for the hydrophilic neutral OMPs in the presence of membrane fouling. The higher rejection corresponded to higher hydrophilicity of the FO membrane, which is fouled rather than clean membrane. Negatively charged membrane surface caused by foulant composition increased adsorption capacity, thereby decreasing mass transport capacity accordingly⁹⁹. Concern on FO process is that appropriate draw solution with high salt concentration is always needed to provide osmotic pressure to the membrane. Additional process to reconcentrate draw solution also needed in the sustainabule operation of FO membrane. The effect of the foulant on the rejection of OMPs is also reported to be pore size-dependent. The effect of the humic
acid foulant on the rejection of OMPs was investigated by different NF membrane pore sizes
⁹. The enhancement of OMPs rejection through NF270 membranes was associated with the
pore blocking due to hydrophobic interactions between OMPs and membrane⁹. It was found
that about 99 % of the removal efficiency of OMPs was achieved by combining forward
osmosis membrane with AnMBR for the treatment of its effluent ⁹². Transport of OMPs
through RO membrane was influenced significantly by their surface charge. The rejection of
the hydrophobic OMPs was improved owing to hydrophobic-hydrophobic repulsion as
hydrophobic membrane was used to reject them⁹². Another study also showed that membrane
distillation process complements AnMBR to improve the removal efficiency of 26 OMPs
with 76 % of overall removal efficiency ⁹³.

The OMPs which are not degraded biologically can be removed by the absoption of
photon from the wavelength found in sunlight, especially Visible Light, Infrared Radiation
and UV light. Photo-oxidation involves the break-up of OMPs by photons into smaller pieces
and the change of their shape into the form that can be altered by proteins denaturing and the
supplement of other atoms or molecules¹⁰⁰. Recently, there has been attempts to observe the
influence of photo-oxidation to the removal of OMPs from the effluent treated by AnMBR⁸⁹.
Arbagues *et al.* (2013) observed the effect of light to the removal of four different
micropollutants, such as 4-(1,1,3,3-tetramethylbutyl)phenol, technical-nonylphenol (t-NP), 4-
n-nonylphenol (4-NP), and Bisphenol-A (BPA). Although the removal rate varied depending
upon the types of OMPs used as compared to aerobic condition, almost complete removal of
4-NP was observed by AnMBR-Photooxidation process ⁸⁹. Assessing degradation rate of
EDC present in the effluents produced by AnMBR was also studied by using microalgae.
565

High dissolved oxygen concentrations produced by algae can enhance the degradation rates
and as such light intensity should be one of dominant parameters¹⁰¹.

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**Future perspectives and research efforts.**

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This paper describes the state-of-the-art of removal behavior of OMPs observed in AnMBR for improving system designs and operations. Obviously, the AnMBR provides great way to produce high removal efficiency of the OMPs at low energy consumptions from various wastewaters. Although many efforts to better understand the removal mechanisms of the OMPs by AnMBR system have been conducted, more solid understanding of it needs to be achieved. Practical application of the AnMBR to remove OMPs from wastewaters put forth thus far by research community is still very limited. From this review, it is indeed that the removal of OMPs from the AnMBR has good correlation with operational conditions, thus further exploration to investigate these features systematically is desired. Although the extent of OMPs removal efficiency in AnMBR is understood, it varies greatly depending of the nature of the contaminants, far less works has been reported on real wastewaters. Because of fundamental differences among removal pathways, as well as modes of operation and reactor designs, modeling and experimental approaches to understanding dominant removal mechanisms need to be developed. Elucidating metabolic pathways during the biotransformation of OMPs should also need further studies particularly under various temperature regimes. It is always attractive to develop integrated AnMBR process, but more in-depth understanding on the removals of OMPs which may occur in pre- or post-treatment combined with the AnMBR reactor should help to optimize reactor design parameters and intensify the process.

588
Conclusions
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Primary pathway affecting the removals of OMPs in AnMBR is the sorption which is driven by physicochemical characteristics of both contaminants and biomasses present in bulk and/or grown in suspended carriers. Longer SRT can increase the withdrawalss of OMPs due to higher biomass level and microbial diversity. In addition, the sorption into biomass can be improved by increasing HRT. Sorption-desorption between OMPs and biomass should be a rate-limiting step to determine the bio-transformative rate. Although intrinsic removal of the OMPs through the porous membrane often applied in AnMBR is very low, the biofilm formed on membrane surface can improve their removal significantly as much as the Dissolved and Colloidal Matter (DCM) retention. Similarly, biomass carriers such as carbon-based media used in AnMBR should contribute to the enhancement of biosorption and biotransformation of OMPs as operational time is increased. In addition, high temperature (around 40°C), biogas production and sparging significantly increased OMP volatilization according to Arrhenius analysis

604 Acknowledgment

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This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Korea government(MSIT) (2019R1A2C1087530).

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