Enhanced degradation of antibiotics by photo-fenton reactive membrane filtration

Journal of Hazardous Materials 386 (2020) 121955

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Enhanced degradation of antibiotics by photo-fenton reactive membrane filtration

T

Shaobin Suna,b, Hong Yaoa,*, Wanyi Fuc, Shan Xueb,c, Wen Zhangb,c a

Beijing International Scientific and Technological Cooperation Base of Water Pollution Control Techniques for Antibiotics and Resistance Genes, Beijing Key Laboratory of Aqueous Typical Pollutants Control and Water Quality Safeguard, Department of municipal and environmental Engineering, School of civil engineering, Beijing Jiaotong University, Beijing, 100044, PR China b School of Environmental and Municipal Engineering, Qingdao University of Technology, Qingdao, 266033, PR China c John A. Reif, Jr. Department of Civil and Environmental Engineering, New Jersey Institute of Technology, 07102, the US

G R A P H I C A L A B S T R A C T

A R T I C LE I N FO

A B S T R A C T

Editor: Navid B. Saleh

Micropollution such as pharmaceutical residuals potentially compromises water quality and jeopardizes human health. This study evaluated the photo-Fenton ceramic membrane filtration toward the removal of sulfadiazine (SDZ) as a common antibiotic chemical. The batch experiments verified that the photo-Fenton reactions with as Goethite (α-FeOOH) as the photo-Fenton catalyst achieved the degradation rates of 100% within 60 min with an initial SDZ concentration of 12 mg·L−1. Meanwhile, a mineralization rate of over 80% was obtained. In continuous filtration, a negligible removal rate (e.g., 4%) of SDZ was obtained when only filtering the feed solution with uncoated or catalyst-coated membranes. However, under Ultraviolet (UV) irradiation, both the removal rates of SDZ were significantly increased to 70% (no H2O2) and 99% (with H2O2), respectively, confirming the active degradation by the photo-Fenton reactions. The highest apparent quantum yield (AQY) reached up to approximately 25% when the UV254 intensity was 100 μW·cm-2 and H2O2 was 10 mmol·L−1. Moreover, the photo-Fenton reaction was shown to effectively mitigate fouling and prevent flux decline. This study demonstrated synchronization of photo-Fenton reactions and membrane filtration to enhance micropollutant degradation. The findings are also important for rationale design and operation of photo-Fenton or photocatalytic membrane filtration systems.

Keywords: Photo-Fenton Ceramic membrane Antibiotics SDZ α-FeOOH

⁎ Corresponding authors at: Beijing Key Laboratory of Aqueous Typical Pollutants Control and Water Quality Safeguard, Department of municipal and environmental Engineering, School of civil engineering, Beijing Jiaotong University, Beijing, 100044, PR China. E-mail address: [email protected] (H. Yao).

https://doi.org/10.1016/j.jhazmat.2019.121955 Received 24 July 2019; Received in revised form 26 November 2019; Accepted 21 December 2019 Available online 24 December 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.

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1. Introduction

photo-irradiation(Hui et al., 2013; Sun et al., 2018). Catalyst coating not only enables reactive membrane filtration but also help retain catalysts and prevent their aggregation. Furthermore, the filtration process may enhance the convective mass transfer rates of pollutants on the coated catalysts and thus improve the reaction kinetics compared to reactors with suspended catalysts (Fu and Zhang, 2018). The surfaceenhanced reactions could lead to improved antifouling or defouling characteristics during membrane filtration as demonstrated previously (De Angelis and de Cortalezzi, 2016; Sun et al., 2018; Yu et al., 2017). Besides the improved antifouling capabilities, reactive membrane filtration could also be effective toward the degradation of refractory water contaminants, which however has not been extensively investigated. This study evaluated the utility of the photo-Fenton ceramic membrane filtration toward the removal and degradation of SDZ as a model pharmaceutical residual micropollutant. Goethite (α-FeOOH) was chosen as the photo-Fenton catalyst due to its excellent photocatalytic activity and stability (Zhang et al., 2016). α-FeOOH catalysts were coated via covalent binding on planar ceramic membranes. Degradation experiments were carried out by filtering the feed SDZ solutions with/ without H2O2 and UV irradiation. We assessed the level of the electron loading from the influent pollutant (Je) and the electron transfer rate (JP) in the photo-Fenton reactions on ceramic membrane surface. This assessment eventually led to the determination of the apparent quantum yield (AQY). Finally, the total organic carbon (TOC) removal and degradation by-products were both assessed under different degradation conditions to achieve deeper insight into the degradation mechanisms during this photo-Fenton reactive membrane filtration.

Unregulated organic micropollutants are increasingly identified in natural waters and commonly recognized as as contaminants of emerging concern (CECs) (Reemtsma et al., 2006). These micropollutants may include pharmaceutical compounds and industrial additives such as flame retardants and organic solvents (Klamerth et al., 2013). Conventional wastewater treatment plants are not designed to effectively remove these micropollutants (Bueno et al., 2012), which may end up in the treatment plant effluents and compromise the quality of receiving waters (Escher et al., 2011; Kasprzyk-Hordern et al., 2009; Pal et al., 2010). Antibiotics, among other CECs, are one of the widely used pharmaceuticals that prevent bacterial infections on human and veterinary animals with the largest global application or consumption (Perini et al., 2018). In recent years, antibiotics are frequently detected in municipal and surface waters due to the ineffective treatment or improper disposal from consumers (Kolpin et al., 2002; Phillips et al., 2010). For example, sulfadiazine (SDZ), a typical antibiotics compound, has been found in aquatic environments ranging from 0.04 ng L−1 to 5.15 g L−1 (García-Galán et al., 2009; Jesús et al., 2010). SDZ is frequently used to treat urinary tract infection, especially in livestock. SDZ enters the environment via different routes, particularly by surface runoff from contaminated soil or improperly disposed animal feces (Boxall et al., 2003). Moreover, Verlicch et al. reported that in hospital discharges antibiotics that could be 2–150 times higher than the concentrations in municipal wastewater (Verlicchi et al., 2010). Although most of the detected antibiotics in the environment are at low concentrations (e.g., 110–610 ng L−1) (Guo et al., 2018), concerns about potential bioaccumulation and persistent environmental impacts are incresaing (Clarizia et al., 2017). Advanced oxidation processes (AOPs) such as UV-based photocatalysis, ozonation, electrochemical oxidation, Fenton and Fenton-like processes are often used to treat and mineralize diverse recalcitrant organic pollutants (Asghar et al., 2015; Chong et al., 2012; Tao et al., 2018). Particularly, there has been a rapid growth of interests in the development of catalytically reactive membrane filtration or reactors to tackle the challenges in the micropollutant removal (Qing et al., 2019). Typical challenges for regular physical filtration include the low rejection efficiency toward micropollutants with small sizes or molecular weights and membrane surface fouling (Kumari et al., 2020). For instance, despite that reverse osmosis (RO) process removed up to 95–99% contaminants in water, the residual Ibuprofen and diclofenac in permeate water could reach a level of 28–223 ng·L−1, implying that RO does not act as an absolute barrier for antibiotics (Sahar et al., 2011). The performances of catalytic membranes such as selectivity, rejection, and antifouling capability could be tailored by anchoring of functional catalysts to the membrane material via surface coating, doping, immobilization and surface grafting (Molinari et al., 2017). Different semiconductor catalysts such as TiO2, SnO2, ZrO2 and WO3 have been used in fabrication of photocatalytic membranes (Chong et al., 2010; Darowna et al., 2014; Ramasundaram et al., 2013). Photocatalysts have been immobilized on a broad range of membranes such as alumina or titania ceramics or polymers such as poly(vinylidene fluoride) (PVDF), poly(sulfone) (PSF), poly(ethersulfone) (PES), poly (acrylonitrile) (PAN), and even metallic substrate such as stainless steel. For example, stainless steel membranes were modified with nanoscale coatings of graphene to promote electrochemical degradation of paracetamol (Huong Le et al., 2019). In our previous work (Sun et al., 2018) and a few others (De Angelis and de Cortalezzi, 2016), heterogeneous photo-Fenton processes are coupled with membrane filtration to enhance chemical degradation (Karnik et al., 2005; Schlichter et al., 2004), filtration performance (Li et al., 2009) and also antifouling or self-cleaning capabilities (Al-Attabi et al., 2019; Guo et al., 2016b; Kim et al., 2008; You et al., 2007). For example, the ferrioxalate-mediated photo-Fenton process reduces Fe(III) ions through ligand–metal charge transfer and activates H2O2 to produce hydroxyl radicals (•OH) under

2. Material and method 2.1. Functionalization of ceramic membrane Synthesis of α-FeOOH catalysts and surface coating on the flat-sheet ceramic membrane with the average pore size 0.14 μm (Sterlitech Corporation, US) were conducted following the reported methods (Sun et al., 2018). Briefly, α-FeOOH catalysts were synthesized in a precipitation method, where 0.5 mol·L−1 Fe(NO3)3 was titrated with a 2.5 mol·L−1 NaOH solution until the reaction solution reached a pH of 12. Then, the suspension was oven dried at 60 °C for 12 h and cooled at room temperature. The precipitate washed repeatedly with DI water and vacuum dried at 60 °C for 2 h. The synthesized catalyst particles with a length of approximately 400–500 nm and the width of about 25–50 nm were immobilized onto planar membranes using Bis-(3[triethoxysilyl]-propyl)-tetrasulfide (22.3%, w/w, in water) as a silane binder. The characterization of the synthesized catalysts and catalyst-coated membranes was reported in our previous work (Sun et al., 2018), including morphology, crystallinity, permeability, and pore sizes (before/ after catalyst coating) using a suite of techniques (e.g., SEM-EDX, XRD, and FTIR). To verify the structural and chemical properties of the catalyst-coated membranes before/after filtration experiments in this study, the catalyst surface coverage and thickness were examined by Scanning electron microscopy (SEM, FEI QUANTA 200). Raman microscope (Renishaw inVia Raman Microscope, UK) using laser power of 5 mW with a 633 nm laser excitation at room temperature was employed to acquire Raman spectra for the origin and coated membrane. 2.2. Batch degradation experiments under different conditions To examine the contributions of membrane adsorption, UV photolysis and photo-Fenton reaction toward the removal of SDZ, a series of bath experiments were carried out with/without UV254 irradiation, H2O2 and the presence of α-FeOOH catalyst on the membrane. Briefly, 30 mL of the SDZ solution with the initial concentration of 12 mg·L−1 was prepared. Then, the ceramic membrane (47 mm in diameter and 2

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filtration tests with the constant initial SDZ concentration 12 mg·L−1 unless indicated otherwise.

2.5 mm in thickness) with or without the coating of α-FeOOH was placed on the bottom of the 90-mm petri dish as shown in Fig. S1 in the supporting information (SI). The α-FeOOH catalysts were coated on ceramic membranes using bis-(3-[triethoxysilyl]-propyl)-tetrasulfide as a covalent binder (Sun et al., 2018). The coating amount was varied by placing a proper volume of the catalyst-binding mixture evenly on ceramic membrane to achieve approximately 2 μg-catalyst g−1-membrane, which was optimized in our previous work. The distance between the UV lamp and the surface of the liquid was 2.5 cm to obtain approximately the exposure intensity of 401 μW·cm-2. The petri dish was mildly agitated on a rotational shaker to thoroughly mix up the solution. The dose of H2O2 was consistently 10 mM for all the experiments unless indicated. 0.5 mL samples were taken at different times (0, 1, 5, 10, 20, 30, 60 min) and filtered before the analytical measurement of the SDZ concentrations by a high-performance liquid chromatography (HPLC, WATERS e2695, USA) as detailed in the SI.

2.3.3. Fouling assessment and flux recovery The fouling and defouling processes of the membrane were indicated by the changes of trans-membrane pressures (TMP). Hydraulic pressure in the feed stream (PF) was monitored by a pressure gauge (PEM-LF SERIES, WINTERS). The hydraulic pressure in permeate stream (Pp) was open to the atmosphere. Thus, the TMP is equal to (PF − PP ) . In our study, the membrane is considered significantly fouled and deserves membrane backwash when the TMP increase by 50% of the initially applied TMP (∼4.2 psi). Thus, the fouling assessment experiments were conducted till significant fouling occurred. Then, the membrane was backwashed by DI water for 30 min to remove the retained pollutants on membrane surface or pores. Alternatively, the fouled ceramic membrane was chemically washed with a 15–20 g·L−1 NaOH solution at 80 °C for 30 min and followed by cleaning with 75% phosphoric acid (H3PO4) at 50 °C for 15 min. After chemical cleaning, the ceramic membrane was raised with DI water repeatedly to remove residual chemicals.

2.3. Filtration experiments 2.3.1. Operation of continuous filtration experiments The removal and degradation of SDZ were also assessed in a deadend mode filtration through the catalyst-coated ceramic membrane. The membrane filtration module was made of polytetrafluoroethylene (PTFE) that is highly resistant to chemical oxidation or UV irradiation. The available membrane surface area was approximately 17.34 cm2 with an overhead space of 1.9 ml (0.2 cm in depth) and a quartz window allowing the UV light illumination (Fig. S2). A UVL 214-Watt lamp (Analytikjena Company, US) provides a monochromatic UV254 irradiation of 401 μW·cm−2 on the surface of the α-FeOOH-coated membrane.

2.4. Analysis of photocatalytic degradation mechanisms The concentrations of hydroxyl radical (%OH) under different reactions were determined versus time using p-Chlorobenzoic acid (pCBA) as the probe indicator of %OH (Li et al., 2012). pCBA purchased from Sigma-Aldrich was prepared and added to the reaction solution at 20 mg L−1, where pCBA reacted with the photogenerated %OH. Under different reaction times, aqueous samples were taken to measure the concentration of remaining pCBA (See details in Section S1). The temporal evolution of the UV–vis spectra of the SDZ reaction solution was detected by the TU-1900 spectrophotometer from 200 nm to 800 nm to reveal the degradation process of SDZ. The TOC levels in the reaction solution was measured by the ELEMENTAR VARIO TOC analyzer (Elementar Analysensysteme, Germany). To examine the photocatalytic degradation pathways for SDZ, influent and effluent samples from the photo-Fenton reactions were analyzed for degradation byproducts using liquid chromatographyelectrospray ionization mass spectrometry equipped with an electrospray ionization source (ESI) or LC-ESI-MS (Agilent1290-6430, USA).

2.3.2. Pollutant (electron) loading and flux control The feed solution was pumped into the filtration module at a constant flow using a LEADFLUID pump (BT100 L). To select proper levels of influent flow, we hypothesize that the level of the electron loading from the influent pollutant (Je) should be equal or less than the maximum possible capacity of electron transfer rate (JP) of the photoFenton reactions on ceramic membrane surface. The electron transfer rate could be affected by the applied UV irradiation intensity, available reaction sites, the photogenerated electron-hole separation and other factors (e.g., catalyst types or H2O2). The rate (e−·s-1) of electrons (JP) transferred from valence band to the conduction band of the photocatalyst can be calculated by: JP =η × UV intensity × surface area/band gap

2.5. Statistical analysis The following experiments were carried out with triplicate independent sampling or testing: (1) Degradation assessment of SDZ in batch mode (Figs. 1 and 2); (2) Hydroxyl radical measurement (Fig. 5); (3) The concentration measurement of SDZ and TOC. The presented results are usually presented with average values with standard deviation as error bars. For the filtration studies, permeate samples were taken at multiple sampling times to obtain stable and representative results, which were shown as average (Figs. 3 and 4). For the fouling/ defouling assessment, only one typical set of experimental data (e.g., TMP vs time) was shown in Fig. 7. However, three repetitions of filtration tests were conducted to confirm the observations. t-test was used to examine the significance of data variations we observed in Fig. 3 in different filtration conditions using at a significant level of 0.05.

(1)

where η is the apparent quantum yield (e.g., 5–15 %), the UV intensity was 401 μW·cm−2, the effective UV-exposure area was about 12.56 cm2 and the band gap of FeOOH catalyst was 2.5 eV (1 eV = 1.6 × 10-19 J). Therefore, under the current experimental conditions, the maximum rate of electron transfer (JP) is 1.26 ± 0.63 × 1015 e−·s−1, assuming η = 15%. The electron loading rate (Je) from the influent pollutants (i.e., SDZ) is a function of the flow rate (Q) and concentration of SDZ:

Je = nCsdz

(2) −1

−1

where CSDZ is the concentration of SDZ (mg·L or mol·L ), n is the number of the total electrons in each SDZ molecule (#·mole−1), which varies from 30 to 38 depending on the oxidation degree (See detailed analysis in the SI). To achieve Je ≤ JP , the influent flow rate or flux should be controlled as shown in Fig. S3. If there is a high n value or a high initial SDZ concentration, the influent flow rate (Q) should be reduced. When the initial SDZ concentration is fixed, a greater quantum yield (η) would increase the allowable maximum flow rate of the SDZ feed solution. In our experiments, we consistently selected 5 μL·s-1 or 3 × 10−4 L·min−1 (influent flux: ∼14.3 LMH) in the dead-end

3. Results and discussions 3.1. Assessment of catalyst integrity before and after filtration studies Our SEM images in Fig. 1 show that the surface coating on the membrane remained unchanged or at least visibly intact before and after filtration of the solution of 20 mM H2O2. A batch experiment was also conducted to confirm that the FeOOH did not dissolve in contact with H2O2. The coated membrane was immersed in the 20-mM H2O2 3

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Fig. 2. The ratio of remaining concentration (C) of SDZ over the initial concentration (C0) under different degradation processes on catalyst-coated ceramic membrane. Initial SDZ concentration: 12 mg·L−1. The UV wavelength was 254 nm and the UV intensity was 401 μW·cm−2; H2O2 concentration was 10 mmol·L−1; and the catalyst on the ceramic membrane was 2 μg·g-1.

Fig. 1. SEM images of the coated membrane (a) before and after (b) filtration of 20 mM H2O2 in water (TMP: 5 psi; filtration time: 12 h). (c) Raman spectra of the α-FeOOH catalyst and the catalyst-coated ceramic membranes.

solution for 12 h. The weight of the treated membrane was compared before and after the immersion had negligible changes. Thus, the dissolution of FeOOH catalyst in contact with H2O2 was not significant. Particularly, the 20 mM H2O2 solution renders a neural pH of 6.51 and a redox level of 211 mV, which does not promote the dissolution of FeOOH (Samson and Eggleston, 2019). Iron dissolution occurs primarily due to (i) reductive dissolution when structural iron dissolves as Fe2+ (i.e., ferrous iron) species and (ii) acidic dissolution when it dissolves as Fe3+. For instance, the dissolution of structural iron is facilitated by protonation of surface oxygen atoms at acidic pH (< 7). These protonation reactions can operate alone or in concert with reduction, further destabilizing the Fe − O bonds and lowering the activation barrier for iron dissolution (Klyukin et al., 2018). The Raman spectrum of α-FeOOH catalyst and catalyst-coated

Fig. 3. (a) The degradation kinetics of SDZ and (b) The apparent quantum yield (QY) under different UV254 irradiation intensities. Initial SDZ concentration: 12 mg·L−1; H2O2 concentration was 10 mmol·L−1; and the catalyst on the ceramic membrane was 2 μg·g−1.

membrane exhibited two prominent peaks: the T peak at ∼295 cm−1 and E peak at ∼403 cm−1, which match the standard peaks of αFeOOH catalyst (Kustova et al., 1992). However, the two Raman peaks on catalyst-coated membrane and the coated membrane slightly shifted 4

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indicative of no substantial chemical changes on surface catalyst. 3.2. Assessment of pollutant degradation in batch experiments 3.2.1. Photocatalytic degradation performance of catalyst-coated ceramic membranes Fig. 2 compares the degradation rates of SDZ on catalyst-coated membranes under different conditions. The removal rate of SDZ was negligible when the SDZ solution was only exposed to catalyst-coated membrane, implying that the surface adsorption of SDZ on coated membrane was minor. Similarly, the SDZ degradation was also negligible if only H2O2 was present in the solution. When the membrane was present with addition of H2O2, the SDZ removal slightly increased to a stable level of over 20% after 10 min. By contrast, the SDZ removal was significantly improved under UV irradiation, which alone led to a progressive SDZ degradation as shown by the purple triangle data. With the combination with UV/H2O2 or the coated membrane/UV/H2O2, the SDZ removal efficiencies were substantially increased. UV irradiation alone appeared to cause SDZ degradation or photolysis, especially in the presence of the catalyst-coated membranes, on which UV photocatalytic reactions may occur. The results in Fig. 2 were fitted using a first-order degradation kinetics (Yang et al., 2015). The corresponding rate constants (k) and the squared correlation coefficients (R2) are summarized in Table S1. The highest reaction rate constant was obtained when using the coated membrane under UV/H2O2, confirming that photo-Fenton reaction on the membrane was the primary factor for the enhanced degradation of SDZ (Yadav et al., 2018).

Fig. 4. The removal of SDZ under different filtration conditions: SDZ concentration: 12 mg·L−1; Influent flux: 10 LMH; UV intensity: 401 μW·cm−2; H2O2 dosage: 10 mmol·L−1 at 5 ± 0.2 μL·s-1 and CM denotes for coated membrane. All error bars represent standard deviation from triplicate experiments. The differences between the control data (uncoated membrane) and other four test data groups were examined for significance using t-test at a significant level of 0.05. * indicates no significant difference (p > 0.05).

3.2.2. Effect of UV intensity on the SDZ photodegradation Fig. 3 shows the degradation kinetics of SDZ at different UV intensities from 100 to 400 μW·cm−2. The degradation rate of SDZ increased monotonously with the increase of the UV intensity, suggesting that the degradation of SDZ was primarily controlled by photo irradiation. However, when the UV intensity increased from 200 to 300 μW·cm−2, the degradation rate did not increase as significantly as expected as the degradation kinetics become limited by the reaction sites, whereas the UV irradiation may have reached the absorption capacity of the coated catalysts. To verify this speculation, we calculated the apparent quantum yield (AQY) for the SDZ degradation under four different levels of UV irradiation:

AQY =

#of SDZ degraded per time #of UV photons per time

(3)

Since the degradation rate varied with the reaction time, the stable degradation rates around 5 min were used in the calculation. The UV photon flux (#photon·cm−2·s-1) under different UV intensities (μW·cm−2) are computed by the Einstein's equation and then converted to the rate of applied photons (#photon·s-1) by multiplying the UV-irradiated area as mentioned above. Fig. 3b shows that the highest AQY was approximately 25%, highlighting the high efficiency of photoFenton degradation reactions. The AQY gradually declined when increasing the UV intensity from 100 to 400 μW·cm−2. This result indicates that the UV irradiation applied to the membrane surface should be optimized based on the available reaction sites of the coated catalyst. This result also implies that the applied UV intensity at 200 μW·cm−2 or less in our experiment should reach a reasonably high AQY. Fig. 5. (a) The SDZ removal ratios and electron loading rates at different initial SDZ concentrations with an influent flux of 10 LMH. (b) The removal of SDZ under original and coated ceramic membranes: SDZ concentration: 12 mg·L−1. UV intensity: 401 μW·cm−2 and H2O2 dosage: 10 mmol·L−1 at 5 ± 0.2 μL·s-1.

3.2.3. Effect of H2O2 dose on SDZ photodegradation Fig. S4 shows increasing the concentration of H2O2 could increase the photodegradation rate of SDZ on the coated membrane. Similar to UV irradiation, after H2O2 increased from 10 to 20 mM, the increase of degradation rates became less significant, probably because of the scavenging effect of excessive H2O2 through this reaction: H2O2 + %OH → H2O + HO2%, which sequesters %OH and thus inhibits the photodegradation. The produced hydroperoxyl radicals (HO2%), however, are substantially less reactive than %OH and do not contribute to the

to the D peak at ∼306 cm−1 and G peak at ∼421 cm−1, probably due to the interactions between the catalyst and the ceramic membrane (Bayati et al., 2019). Particularly, the coated membranes before and after filtration of the H2O2 solution resulted in similar Raman spectra, 5

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oxidation of pollutants.(Neamu et al., 2004) Therefore, optimization of H2O2 doses is critical for the photo-Fenton reactions and pollutant degradation. 3.3. Pollutant removal and degradation in continuous filtration 3.3.1. The removal of SDZ under different membrane filtration conditions Fig. 4 shows that approximately 4% of SDZ was removed by the uncoated membrane with less than 1% of TOC reduction, which indicates that the contribution from the size exclusion or membrane surface adsorption was negligible. Meanwhile, filtration through the catalyst-coated membrane slightly increased the removal rate of SDZ to about 10%. However, almost no TOC reduction was achieved. By contrast, when the catalyst-coated membrane was exposed to H2O2, 22% of SDZ was removed with 5% of TOC reduction, indicating that the degradation of SDZ was slightly enhanced but the mineralization was still minor. In the presence of catalyst on the membrane and UV irradiation, both the removal rates of SDZ and TOC were significantly increased to 70% and 40% respectively, which agrees with the results from the batch experiments. Furthermore, when applying UV irradiation and H2O2 onto the catalyst-coated membrane, the removal of SDZ reached the highest level (almost 99%), confirming that the degradation/mineralization was primarily attributed to the photo-Fenton reactions. 3.3.2. The effect of the initial SDZ concentration and influent flux on the SDZ removal Fig. 5a shows that the removal ratios of SDZ in the filtered permeate began to decline when the initial concentration of SDZ was over 24 mg·L−1. This decline was due to the incomplete photocatalytic degradation while filtration as verified by a comparison of the electron loading rate (Je) from pollutant oxidation reactions and the electron transfer rate (JP) in the photo-Fenton reaction. According to Eq. 1 and 2, Je and JP are computed and shown in Fig. 5a, respectively, assuming the apparent quantum yield (η) of 15% and n of 40 e−·mole-1-SDZ. Clearly, when the SDZ concentration was over 24 mg·L−1, JP became lower than Je, which means the electron loading rate from the pollutant flux exceeded the capacity of electron transfer rate (red dotted line) or degradation rate in the current photo-Fenton membrane. Accordingly, the removal rate of SDZ began to decline. Similar to the impact of the initial SDZ concentration, increasing the influent flux essentially increases the level of Je and therefore decreases the degradation rate of SDZ as Fig. 5b demonstrates for both uncoated and coated membranes. The red dotted line points to a level of influent flux that corresponds to the critical flux (16 LMH) when JP=Je (assuming n = 40 e−·mole-SDZ degraded; η = 15%). Based on our results in Fig. 3b, the actual AQY could be up to 25%. Thus, the influent flux (10 LMH) was significantly less than the critical flux and thus, the removal efficiencies of SDZ could reach up to 99% during the filtration with concurrent photo-Fenton reactions. Thus, membrane technologies coupled with photocatalysis yield many advantages over conventional photocatalytic reactors such as in simultaneous separation of water/ photocatalysts and pollutant degradation (Fischer et al., 2014; Kumari et al., 2020).

Fig. 6. (a) Concentration of the produced hydroxyl radical (b) via time under different photo-Fenton or photocatalytic reactions. (b) Temporal evolution of the UV–vis spectra during the photodegradation of SDZ under the photocatalytic condition as shown below. (c) The TOC removal in batch photo-Fenton reactions with or without the presence of the catalyst-coated ceramic membrane. Initial SDZ concentration: 12 mg·L−1 corresponding to an initial TOC concentration of 5.8 mg·L−1, UV wavelength was 254 nm and intensity was 401 μw·cm−2; H2O2 concentration was 10 mmol·L−1, and the catalyst on the ceramic membrane was 2 μg·g-1.

3.4. Analysis of photocatalytic degradation mechanisms The surface sites (≡FeⅢ(OH)) on α-FeOOH catalyst are considered to catalyze the generation of hydroxyl radicals and peroxide anions via photo-Fenton reactions.(Sun et al., 2018) Fig. 6a shows the evolution concentrations of hydroxyl radical under different reaction conditions, whereas Fig. S5 shows the corresponding decline of the pCBA concentrations. As reported previously, (Li et al., 2012) %OH stoichiometrically reacts with pCBA in a mole ratio of 1:1, and degradation of %OH through side reactions with other potential contaminants in our reaction system can be ignored. Our results show that UV or H2O2 alone did

not decrease the pCBA concentration significantly, which similarly occurred to coated membrane with or without the addition of H2O2. When exposed to UV alone or the catalyst-coated membrane, the pCBA concentration started to decline much faster. Furthermore, when UV/H2O2 was applied with or with the catalyst-coated membrane, the pCBA concentration declined sharply, indicative of the generation of •OH via photocatalysis or photo-Fenton reactions. 6

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Fig. 7. Reaction pathways of SDZ and the major intermediates observed in this study.

antibiotics activity of SDZ’s oxidation by-products depends on factors such as oxidation mechanisms and test organisms (e.g., bacteria, fish, daphnid and algae), which deserves extensive and standardized toxicity assessment in future research work.

Fig. 6b shows the temporal evolution of the UV–vis spectra was acquired from the reaction solution. The characteristics absorbance peak at 273 nm indicates the presence of SDZ in the solution. The intensity of this peak decreased over reaction time from 0 to 60 minutes, which matches the degradation kinetics data (C/C0) presented in Fig. 2. This result confirmed the sustained degradation of SDZ under photoFenton reaction on the coated membrane surface. Fig. 6c shows the TOC changes in the SDZ solution under different reaction conditions. No mineralization of SDZ was obtained when the solution was only exposed to the catalyst-coated membrane or H2O2. By contrast, a TOC removal rate of 90% at 60 min when UV/ H2O2 were both applied to the catalyst-coated membrane. However, the mineralization of SDZ was reduced to 49% and 42% at 20 min if the solution was only exposed to UV/H2O2 or to the catalyst-coated membrane under UV irradiation only.

3.6. Fouling assessment on during photo-Fenton reactive membrane filtration Fig. 8a compares the TMP changes during filtration under different conditions. The TMP remained stable around 4.2 psi when filtering the DI water. By contrast, the TMP increased quickly when filtering the SDZ solution (red dots). Based on the molecular size of SDZ (250 Da) and the average membrane pore size (140 nm), the membrane fouling may follow the standard fouling law, where the inner membrane pore size may reduce due to the deposition of SDZ on the pore surface. Adding H2O2 into the influent flow did not slow down membrane fouling, and in instead increased the membrane fouling kinetics (green triangle), which was also observed in our previous study (Sun et al., 2018). The enhanced fouling kinetics is probably caused by complex interactions between SDZ, the membrane surface and H2O2, which may promote the surface deposition and pore blocking by SDZ (Lee et al., 2017; Monash and Pugazhenthi, 2011). When exposed to UV irradiation only, the membrane fouling was substantially mitigated (blue triangle). Combing UV and H2O2 further slowed down the increase of TMP, which demonstrates the antifouling capability of the photo-Fenton reactive membrane. Fig. 8b compares the changes of TMP when filtering the feed solution with various initial concentrations (6, 12, 24 and 48 mg L−1) of SDZ. Under the SDZ level of 24 mg L−1 or less, there was no significant increase of TMP or membrane fouling as the photo-Fenton reaction sufficiently degraded the influent flux of SDZ as we discussed above. However, when the SDZ concentration was greater than 24 mg L−1 (especially 48 mg L−1), the TMP progressively increased as the surface foulant buildup, which occurred when the SDZ flux or the electron loading rate exceeded the capacity of electron transfer via photo-Fenton reactions on the membrane.

3.5. Photocatalytic degradation pathway of SDZ To gain a better understanding of mineralization mechanisms, the oxidation byproducts of SDZ were identified by LC-ESI-MS. Fig. S6 shows the chromatograms of the treated permeate samples at different reaction times (0, 5, 30 and 60 min). Four main oxidation intermediates and SDZ are identified based on their characteristic peaks at different m/z values. Fig. 7 illustrated the hypothetical degradation or transformation pathways (A, B, C and D) of SDZ in the photo-Fenton oxidation reaction, including the pathway A and C: cleavage of the SeN bond, which leads to the hydrolyzation of SDZ into 4-aminobenzenesulfonic acid (m/z = 197) and 2-nitropyrimidine (m/z = 156), respectively. In pathway B, SO2 is eliminated with the formation of 4-(2-iminopyrimidin-1(2 H)-yl)　aniline (m/z = 215). In pathway D, NH2 in the benzene ring was oxidized, perhaps by radicals produced in the photoFenton reactions. Probably because the concentrations of these intermediates and byproducts were too low to be detected (lower than the detection limits of GC or LC), all prior studies on the catalytic degradation of SDZ did not determine the byproduct concentrations (Acosta-Rangel et al., 2018; Feng et al., 2016; Guo et al., 2016a; Ji-Feng, 2014; Liu et al., 2009; Rong et al., 2014; Xuan and Huong, 2016; Wang et al., 2019; Yadav et al., 2018; Yang et al., 2018). Additionally, most byproducts do not have commercially available standard chemicals that are needed for quantification. Therefore, the toxicity changes of treated water were usually assessed using the treated water as a mixture instead of attributing to any of the single byproducts. Some research suggested the toxicity or antibiotics activity could be reduced in the treated SDZ water by a Fenton-like Fe3+/calcium peroxide (CaO2) reaction (Lu et al., 2020), potassium persulphate (Gokulakrishnan et al., 2012) or Ferrate (VI) (Wang et al., 2019), while other studies using permanganate (Yang et al., 2018), chlorination (Dong et al., 2019) and Fenton (Ji-Feng 2014) reported the formation of byproducts that exhibited higher toxicity than SDZ. Apparently, the remaining toxicity or

4. Conclusion In this study, the photo-Fenton reaction was achieved on a porous ceramic membrane as a support, on which micropollutant degradation reactions and membrane filtration were synchronized. The removal of sulfadiazine (SDZ) was enhanced in the process of photocatalytic membrane filtration compared to the batch mode reaction, which may be attributed to the forced convection and improved mass transport of reactants on the immobilized catalysts. Besides the enhanced pollutant degradation, this photocatalytically reactive membrane exhibited strong antifouling capability that effectively sustained longer filtration time. The catalyst coated on ceramic membranes were stable during the filtration experiments. This study further compared the pollutant or 7

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contributed to the final manuscript. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgement This study was supported by the China’s National Science Foundation (Award Number: 51778306 and 51578043) and Fundamental Research Funds for the Central Universities (Award Number: 2016JBZ008), Beijing Outstanding Young Scientist Program (BJJWZYJH01201910004016) and the Beijing Jiaotong University fundamental Funds (Award Number: 2018YJS307). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.121955. References Acosta-Rangel, A., Sánchez-Polo, M., Polo, A.M.S., Rivera-Utrilla, J., Berber-Mendoza, M.S., 2018. Sulfonamides degradation assisted by UV, UV/H2O2 and UV/K2S2O8: Efficiency, mechanism and byproducts cytotoxicity. J. Environ. Manage. 225, 224–231. Al-Attabi, R., Rodriguez-Andres, J., Schütz, J.A., Bechelany, M., des Ligneris, E., Chen, X., Kong, L., Morsi, Y.S., Dumée, L.F., 2019. Catalytic electrospun nano-composite membranes for virus capture and remediation. Sep. Purif. Technol. 229, 115806. Asghar, A., Abdul Raman, A.A., Wan Daud, W.M.A., 2015. Advanced oxidation processes for in-situ production of hydrogen peroxide/hydroxyl radical for textile wastewater treatment: a review. J. Clean. Prod. 87 (Supplement C), 826–838. Bayati, M., Peng, H., Deng, H., Lin, J., Fidalgo de Cortalezzi, M., 2019. Laser induced graphene /ceramic membrane composite: preparation and characterization. J. Memb. Sci., 117537. Boxall, A.B.A., Fogg, L.A., Kay, P., Blackwel1, P.A., Pemberton, E.J., Croxford, A., 2003. Prioritisation of veterinary medicines in the UK environment. Toxicol. Lett. 142 (3). Bueno, M.J.M., Gomez, M.J., Herrera, S., Hernando, M.D., Agüera, A., Fernández-Alba, A.R., 2012. Occurrence and persistence of organic emerging contaminants and priority pollutants in five sewage treatment plants of Spain: two years pilot survey monitoring. Environ. Pollut. 164 (Supplement C), 267–273. Chong, M.N., Jin, B., Chow, C.W.K., Saint, C., 2010. Recent developments in photocatalytic water treatment technology: a review. Water Res. 44 (10), 2997–3027. Chong, M.N., Sharma, A.K., Burn, S., Saint, C.P., 2012. Feasibility study on the application of advanced oxidation technologies for decentralised wastewater treatment. J. Clean. Prod. 35 (Supplement C), 230–238. Clarizia, L., Russo, D., Di Somma, I., Marotta, R., Andreozzi, R., 2017. Homogeneous photo-Fenton processes at near neutral pH: a review. Appl. Catal. B 209, 358–371. Darowna, D., Grondzewska, S., Morawski, A.W., Mozia, S., 2014. Removal of non-steroidal anti-inflammatory drugs from primary and secondary effluents in a photocatalytic membrane reactor. J. Chem. Technol. Biotechnol. 89 (8), 1265–1273. De Angelis, L., de Cortalezzi, M.M.F., 2016. Improved membrane flux recovery by Fentontype reactions. J. Memb. Sci. 500, 255–264. Dong, F., Li, C., Crittenden, J., Zhang, T., Lin, Q., He, G., Zhang, W., Luo, J., 2019. Sulfadiazine destruction by chlorination in a pilot-scale water distribution system: kinetics, pathway, and bacterial community structure. J. Hazard. Mater. 366, 88–97. Escher, B.I., Baumgartner, R., Koller, M., Treyer, K., Lienert, J., McArdell, C.S., 2011. Environmental toxicology and risk assessment of pharmaceuticals from hospital wastewater. Water Res. 45 (1), 75–92. Feng, Y., Wu, D., Liao, C., Deng, Y., Zhang, T., Shih, K., 2016. Red mud powders as lowcost and efficient catalysts for persulfate activation: pathways and reusability of mineralizing sulfadiazine. Sep. Purif. Technol. 167, 136–145. Fischer, K., Gläser, R., Schulze, A., 2014. Nanoneedle and nanotubular titanium dioxide – PES mixed matrix membrane for photocatalysis. Appl. Catal. B 160-161, 456–464. Fu, W., Zhang, W., 2018. Microwave-enhanced membrane filtration for water treatment. J. Memb. Sci. 568, 97–104. García-Galán, Ma.J., Garrido, T., Fraile, J., Ginebreda, A., Díaz-Cruz, M.S., Barceló, D., 2009. Simultaneous occurrence of nitrates and sulfonamide antibiotics in two ground water bodies of Catalonia (Spain). J. Hydrol. (Amst) 383 (1). Gokulakrishnan, S., Parakh, P., Prakash, H., 2012. Degradation of Malachite green by Potassium persulphate, its enhancement by 1,8-dimethyl-1,3,6,8,10,13-hexaazacyclotetradecane nickel(II) perchlorate complex, and removal of antibacterial activity. J. Hazard. Mater. 213-214, 19–27. Guo, J., Farid, M.U., Lee, E.-J., Yan, D.Y.-S., Jeong, S., Kyoungjin An, A., 2018. Fouling behavior of negatively charged PVDF membrane in membrane distillation for removal of antibiotics from wastewater. J. Memb. Sci. 551, 12–19.

Fig. 8. The TMP changes during the filtration of the SDZ solution using the catalyst-coated ceramic membrane solution under conditions: SDZ concentration: 12 mg·L−1. UV intensity: 401 μw·cm−2 and H2O2 dosage: 10 mmol·L−1 at 5 ± 0.2 μL·s-1.

electron loading rates and the transfers of photogenerated electrons to determine the optimal influent flux to the membrane for optimal degradation performances. We also analyzed the roles of hydroxyl radical generation in the degradation mechanisms and unraveled the potential degradation pathways. The treatment efficacy (catalytic activity and membrane permeability) of this reactive membrane filtration process is highly dependent on membrane catalyst coating (density or thickness) and stability, which clearly deserves future research. Author contribution section Wen Zhang and Shaobin Sun conceived of the presented research idea and developed the theory and experimental plans. Shaobin performed the experimental work and received assistance/advices from Hong Yao, Wanyi Fu and Xueshan, who also verified the analytical methods and experimental results. All authors discussed the results and 8

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