Red mud-activated peroxymonosulfate process for the removal of fluoroquinolones in hospital wastewater

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(1)Red mud-activated peroxymonosulfate process for the removal of fluoroquinolones in hospital wastewater Joohyun Kim, Gnougon Nina Coulibaly, Sunho Yoon, Aymen Amine Assadi, Khalil Hanna, Sungjun Bae. To cite this version: Joohyun Kim, Gnougon Nina Coulibaly, Sunho Yoon, Aymen Amine Assadi, Khalil Hanna, et al.. Red mud-activated peroxymonosulfate process for the removal of fluoroquinolones in hospital wastewater. Water Research, IWA Publishing, 2020, 184, pp.116171. �10.1016/j.watres.2020.116171�. �hal02930221�. HAL Id: hal-02930221 https://hal.archives-ouvertes.fr/hal-02930221 Submitted on 11 Sep 2020. HAL is a multi-disciplinary open access archive for the deposit and dissemination of scientific research documents, whether they are published or not. The documents may come from teaching and research institutions in France or abroad, or from public or private research centers.. L’archive ouverte pluridisciplinaire HAL, est destinée au dépôt et à la diffusion de documents scientifiques de niveau recherche, publiés ou non, émanant des établissements d’enseignement et de recherche français ou étrangers, des laboratoires publics ou privés..

(2) Highlights . Red mud (RM) was used as a novel material for peroxymonosulfate (PMS) activation.. . Hydroxylamine significantly enhanced flumequine (FLU) removal by the PMS/RM system.. . Ciprofloxacin (CIP) and FLU were oxidized via ring cleavage, hydroxylation, decarbonylation, and defluorination.. . Phosphate containing in hospital wastewater significantly inhibited the FLU removal.. . Increasing PMS concentration and its sequential addition resulted in complete mineralization of FLU in HW.

(3) Red mud-activated peroxymonosulfate process for the removal of fluoroquinolones in hospital wastewater. Joohyun Kim1,†, Gnougon Nina Coulibaly2,†, Sunho Yoon1, Aymen Amin Assadi2, Khalil Hanna2,3,*,†, Sungjun Bae1*,†. 1. Department of Civil and Environmental Engineering, Konkuk University, 120 Neungdongro, Gwangjin-gu, Seoul 05029, Republic of Korea. 2. Univ. Rennes, Ecole Nationale Supérieure de Chimie de Rennes, CNRS, ISCR-UMR 6226, F-35000 Rennes, France 3. Institut Universitaire de France (IUF), MESRI, 1 rue Descartes, 75231 Paris, France.. † Contributed equally to this work *Co-corresponding author: bsj1003@konkuk.ac.kr *Co-corresponding author: khalil.hanna@ensc-rennes.fr. A revised manuscript submitted to Water Research June, 2020.

(4) Abstract In this study, a novel peroxymonosulfate (PMS) activation method, which combines a solid waste (i.e., red mud, RM) and a reducing agent (i.e., hydroxylamine, HA), for the oxidative degradation of fluoroquinolones (FQs; i.e., flumequine (FLU) and ciprofloxacin (CIP)) in hospital wastewater (HW) was developed. The addition of HA into the PMS/RM suspension significantly enhanced FLU removal, owing to its ability to enhance the Fe(III)/Fe(II) cycle on the RM surface. The results of the quenching experiments suggested the predominance of SO4•− over •OH in the PMS/RM/HA system. Moreover, owing to the greater reactivity between CIP and SO4•−, CIP removal was more effective than FLU removal. Additionally, the liquid chromatography-mass spectroscopy (LC-MS) analysis revealed that the oxidation of CIP and FLU by PMS/RM/HA occurred via sequential and separate processes, involving ring cleavage, hydroxylation, decarbonylation, and defluorination. Surprisingly, the wastewater components exhibited contrasting effects on FLU removal in HW. Natural organic matter, nitrate and sulfate showed a slight impact on the removal performance of FLU, whereas chloride improved the oxidation extent. However, phosphate significantly inhibited the FLU removal because of its competitive binding at the RM surface and its scavenging effect towards SO4•−. This inhibitory effect was overcome by increasing the PMS concentration and its sequential addition, thus guaranteeing successful mineralization of FLU in HW. These results show that the RM/HA system can be utilized to activate PMS for the removal of antibiotics in wastewater.. Keywords: flumequine; ciprofloxacin; hospital wastewater; red mud; hydroxylamine; peroxymonosulfate.

(5) 1. Introduction According to a report published by the World Health Organization in 2017 (World Health Organization, 2017), resistance to antibiotics is presently considered as one of the greatest threats to health, food security, and development globally. Antibiotics, which are commonly used to prevent and treat bacterial infections, are the most prescribed drugs worldwide and frequently encountered in hospital and urban wastewater (Okeke et al., 1999; RodriguezMozaz et al., 2015). Indeed, the presence of penicillins, macrolides, sulfonamides, fluoroquinolones, and tetracyclines in hospital wastewater (HW) has been reported with concentrations up to 15 µg/L (Reemtsma and Jekel, 2006). The amoxicillin concentration measured in HW of a large German hospital was between 28 and 82.7 µg/L, and the ciprofloxacin in concentrations between 3 and 89 µg/L (Reemtsma and Jekel, 2006). Fluoroquinolones (FQs) are the most potent antimicrobial agents used in human and veterinary medicine. Because their removal during wastewater treatment is incomplete and they are continually being released into the environment, they are frequently detected in aquatic systems (Rodriguez-Mozaz et al., 2015; Watkinson et al., 2009). Particularly, the presence of flumequine (FLU) and ciprofloxacin (CIP) in surface water, groundwater, and sediments at concentrations ranging from nanograms to micrograms per liter have been reported (Zhang et al., 2015). To remove such antibiotics from wastewater, advanced oxidation processes (AOPs) have attracted considerable attention recently, owing to their ability to generate reactive oxygen species (ROS) that can efficiently remove refractory and non-biodegradable organic contaminants (Boczkaj and Fernandes, 2017; Deng et al., 2017). Recently, sulfate radical (SO4•−)-based AOPs have become a new focus for wastewater treatment owing to the better selectivity and longer half-life of SO4•− in comparison with the hydroxyl radical ( •OH)..

(6) Reportedly, peroxymonosulfate (PMS) can be activated by UV (Qi et al., 2019), heat (Xu et al., 2016), or a transition metal (Sang et al., 2020). However, its activation in UV and heat systems is associated with disadvantages, such as high cost and high energy consumption (Zou et al., 2013). Additionally, Co 2+/PMS systems demonstrate the best performance with respect to SO4•− generation (Anipsitakis and Dionysiou, 2004); however, the toxicity of cobalt, as well as the resulting secondary pollution associated with its use, stand as a major setback to the application of Co 2+/PMS systems (Deng et al., 2017; Ji et al., 2013). More recently, it has been reported that replacing cobalt with a nontoxic metal, such as Mn 2+ or Fe2+, opens up new avenues for research into the development of more economical and efficient PMS activation techniques (Cao et al., 2019). Furthermore, to activate PMS for the oxidative degradation of organic contaminants, Fe-containing minerals, including pyrite (FeS2; Feng et al., 2018), hematite (α-Fe2O3; Jaafarzadeh et al., 2017), and magnetite (Fe3O4)/akageneite(β-FeOOH) composites (Li et al., 2019), have been employed, indicating that heterogeneous Fe minerals can be effectively used for PMS activation. Red mud (RM), which is a massive byproduct of the Bayer process that is used to manufacture alumina (Hamid et al., 2018), is well known as a mixed metal composite, primarily composed of Fe2O3 in combination with other species, such as Al2O3, SiO2, CaO, and TiO2. Given that it contains CaO, its disposal into the natural environment can cause environmental problems, such as an increase in the pH of the surrounding soil and groundwater (Mymrin et al., 2003). Therefore, many attempts have been made for its reuse as an economical and eco-friendly alternative for environmental remediation technologies (Bhatnagar et al., 2011). Reportedly, it has been used as an adsorbent for dyes, phenols, and heavy metals (Bhatnagar et al., 2011; Sushil and Batra, 2008). However, its ability to enhance the advanced oxidation of antibiotics, while eliminating the emerging contaminants from wastewater has not yet been clarified. Therefore, the primary objective of this study was to.

(7) investigate the ability of RM to enhance the advanced oxidation of antibiotics, while removing antibiotic contamination from wastewater. Additionally, for the first time, its catalytic activity with respect to the removal of antibiotics from hospital wastewater (HW) under mild conditions was examined, and the ability of hydroxylamine (NH2OH, HA), which is a reducing agent, to accelerate ROS generation in the presence of RM was investigated. In this study, RM was characterized using various surface analysis methods, including Xray fluorescence (XRF), Brunauer–Emmett–Teller (BET) adsorption isotherms, X-ray diffraction (XRD), transmission electron microscopy (TEM), X-ray photoelectron spectroscopy (XPS). The FLU removal capacity of the PMS/RM/HA system at an initial pH of 7 was examined, and the effect of different PMS, HA, and RM concentrations on the FLU removal was evaluated. Additionally, the performance of the PMS/RM/HA system was examined in the separated and combined removal of FLU and CIP in HW. Using liquid chromatography-mass spectroscopy (LC-MS), the FLU and CIP oxidation pathways were identified, and the mechanisms of their removal were investigated.. 2. Materials and methods 2.1. Chemicals and materials FLU (C14H12FNO3 ; > 98%), CIP (C17H18FN3O3; > 98%), sodium persulfate (Na2S2O8), PMS (available as Oxone (KHSO5 • 0.5KHSO4 • 0.5K2SO4)), isopropanol (i-PrOH, C3H8O), tertbutanol (t-BuOH, C4H10O), potassium iodide (KI), potassium hydroxide (KOH), sodium chloride (NaCl), sodium nitrate (NaNO3), HA (NH2OH, 50% wt), sodium sulfate (Na2SO4), sodium sulfite (Na2SO3), and disodium phosphate (Na2HPO4) were purchased from Sigma– Aldrich Inc., USA. Sodium dithionite (Na2S2O4) was purchased from PubChem, whereas sodium hydroxide (NaOH) and hydrochloric acid (HCl, 37% extra pure) were obtained from Acros Organics. Standard leonardite humic acid (LHA) was obtained from the International.

(8) Humic Substances Society, and raw RM, which is obtained as a byproduct of the Bayer process, was obtained from the Korea Institute of Geoscience and Mineral Resources at Daejeon, South Korea. The RM was used as supplied, i.e., without further chemical treatment. It was only rinsed with ultra-pure water (UPW). The physical characteristics of RM were identified by various surface analyses such as XRF, TEM, XRD, BET, zeta potential analyzer, XPS. More details for the surface analyses are provided in supplementary data. To prepare standard solutions, UPW with a resistivity of 18.2 MΩ•cm, obtained from a Millipore Milli-Q system was used. HW was provided by the University Hospital Center of Rennes City (CHU Rennes - Hôpital Pontchaillou, France). CHU Rennes covers a broad range of clinical services and health needs and has a big hospitalization capacity (2K+ beds, 150K+ admissions and 500K+ consultations in 2018) in different areas (medicine, surgery, obstetrics, and gynecology). As a result, the wastewater sample investigated in this work represents well the HW in Europe. Its physico-chemical characteristics are presented in Table 1.. 2.2. Oxidation experiments and analytical methods Sorption experiments were conducted in polypropylene tubes (15 mL) within a pH range of 2−10, under aerobic conditions. HCl and NaOH solutions (0.1 M each) were used to adjust the pH, and RM (0.1 g L-1) was added to 5 μM FLU in a final volume of 10 mL. The suspension was stirred for 24 h, and 1.5 mL of the final solution was filtered. Then, its concentration was determined using an Alliance UV controller high performance liquid chromatography (HPLC) system equipped with an autosampler (Waters 717 plus), a C18 column (250 mm × 4.6 mm i.d., 5 μm), and a UV detector (Alliance UV 2489). The FLU and CIP mobile phase, which was set at a flow rate of 1 mL min-1 in the isocratic mode, consisted.

(9) of a mixture of acetonitrile and water (15/85 v/v and 45/55 v/v, respectively) containing 0.1% formic acid. To avoid the influence of UV and visible light, all oxidation experiments were performed at ambient temperature (20 ± 2 °C) under dark conditions. Different amounts of RM were dispersed in a 200 mL aqueous solution with FQ (FLU or CIP) concentrations of 5 µM After thorough mixing, different concentrations of PMS and HA were added to initiate the reaction, and the mixtures were shaken at 300 rpm at a temperature of 20 ± 2 °C and a pH of 7.0 ± 0.2. The pH adjustment was done within 5 minutes after adding RM and PMS by 0.1 M solutions of HCl and NaOH, respectively. For the experiment with HW, the suspended solid was filtered and 5 µM of FQ was spiked into the HW. To identify CIP and FLU intermediates resulting from the oxidation by the PMS/RM/HA system, a Thermo ultimate 3000 HPLC with an LTQ ion trap mass spectrometer (Thermo Electron Corp., USA) equipped with an ESI source that can be operated in both the positive and negative ion modes was used. More details for LC-MS analysis are provided in supplementary data. To perform quenching experiments using t-BuOH and i-PrOH, the desired quenchers were added into the reaction solutions prior to the addition of HA and PMS. At each sampling time, 1.5 mL of the treated solution was collected for the determination of the concentrations of FQs using HPLC. To provide the direct evidence of radical generation in the PMS/RM/HA system , we prepared two different systems using PMS/RM/HA and H2O2/RM/HA at pH 7 and analyzed the radical generation by electron spin resonance spectrometer (ESR, JES-FA200, JEOL). The suspension containing 100 mM 5,5-dimethyl-1pyrroline N-oxide (DMPO) was injected to ESR tube and analyzed at a resonance frequency of 9.45 GHz, microwave power of 0.998 mW, modulation frequency of 100 kHz, field center.

(10) of 340 mT, modulation amplitude of 1.0 G, sweep width of 1.0 G, time constant of 30 ms and sweep time of 30 s. The effects of the co-presence of inorganic (phosphate, sulfate, chloride, and nitrate anions) and organic (LHA) species on FLU degradation were investigated in UPW under the same conditions. An ion chromatography instrument (IC, DIONEX DX-120) equipped with a conductivity detector, and an IonPac AS19 analytical column (4 × 250 mm), with KOH solution as the mobile phase, was used to determine the concentration of the inorganic species. The amount of Fe leached during the reaction was measured using a UV–vis spectrophotometer (GENESYS 10S, Thermo) at 510 nm following the 1,10-phenanthroline method (Tamura et al., 1974). The concentration of residual PMS was measured by the iodometric titration method (Liang et al., 2008). After 6 h of reaction, 0.1 mL of aliquot was filtered and transferred to a quartz cuvette containing 2.9 mL of KI and NaHCO3 solution (600 mM and 70 mM, respectively) for PMS concentration measurement by UV-vis spectrophotometer at 352 nm (Liang et al., 2008). The leaching of other metals after PMS sequential addition was also monitored using inductively coupled plasma-mass spectrometry (ICP-MS, 7800, Agilent, USA). The degradation products of HA (i.e., NO 2- and NO3 -) were quantified using IC. Mineralization, i.e., the removal of total organic carbon (TOC), was investigated using a TOC-meter (Shimadzu TOC-VCSH). All experiments were performed in triplicates, and demonstrated good reproducibility within an average standard deviation of less than 5%..

(11) 3. Results and discussion 3.1. Characterization of RM XRF analysis demonstrated that RM predominantly consisted of iron oxide (37%), alumina (24%), and silica (10%), in combination with other mineral phases, including TiO 2, Na2O, MgO, CaO, K2O, MnO, and P2O5 (Table S1). Figure 1(a), which illustrates the XRD diffractogram of raw RM, indicates the presence of peaks corresponding to hematite, boehmite (γ-AlO(OH)), anatase (TiO2), calcite (CaCO3), and quartz (SiO2). Based on the zeta potential measurements under different pH conditions (2, 4, 6, and 8), the point of zero charge (PZC) of RM (Fig. S1) was 7.3, and its BET surface area determined from the multipoint N2 adsorption isotherm was 23 m2 g-1. Additionally, the characterization of the morphology and structural features of RM using TEM analysis at different optical magnifications revealed that it consisted of nanosized particles (< 100 nm) with different morphologies (Fig. 1(b)) that were aggregated to form bigger particles (Fig. 1(c)).. 3.2. Enhanced removal of FLU using the PMS/RM/HA system The effect of pH variation on FLU adsorption onto the RM surface is illustrated in Fig. S2. As typically encountered in the adsorption of organic anions onto metal oxides, FLU adsorption was high under low pH conditions, and decreased as the pH increased. This observed sorption behavior can be attributed to the combined effect of pH-dependent FLU speciation (pKa = 6.5) and RM surface charge characteristics (PZC = 7.3). Under pH conditions higher than both pKa and PZC, the FLU and RM surfaces were negatively charged, leading to less favorable electrostatic interactions. To determine the most efficient FLU removal system, the kinetics of FLU removal with and without PMS, RM, and HA were investigated (Fig. 2). Assuming that RM promotes the generation of ROS, leading to the degradation of the target compound, and that the.

(12) concentration of these generated species are constant over the reaction time, the FLU removal kinetics can be described using Eq. (1), which is a pseudo-first-order equation.. [. ]( ). [. ]. ,. (1). where kapp, which represents the rate constant (h-1) of the pseudo-first-order reaction, is obtained by considering the linear regression of ln ([FQs]/[FQs]0) vs. time (t). The control experiment using HA alone demonstrated that the addition of 0.05 mM HA did not have any significant effect on the FLU removal kinetics. Furthermore, using 1 mM of PMS alone or in combination with 0.05 mM HA resulted in an FLU removal of up to 40%. The fast removal occurred within 1 h of reaction could be attributed to direct PMS action (Yin et al., 2018) and/or alkaline-activation by NaOH used for pH adjustment at the initial stage of reaction (Wang and Wang, 2018). When 0.05 g L-1 of RM and 0.05 mM HA were used, the FLU removal efficiency was. 30%, showing the initial adsorption capacity of RM. for the FLU removal. Contrarily, after 6 h of reaction in the RM/PMS system, the FLU removal efficiency observed was 75% (kapp = 0.248 h-1, Table S2), which might be caused by sum of FLU adsorption by RM alone and FLU removal by PMS alone. The FLU adsorption– desorption equilibrium was attained after 1 h (see dotted line in Fig. 2), and finally, complete FLU removal was observed after 6 h of reaction using a system obtained by adding 0.05 mM HA to the PMS/RM system. This new system exhibited a kapp value (0.577 h-1) that was 2.3 times higher than that of the PMS/RM-only system, suggesting that HA plays a crucial role in enhancing FLU removal. The enhanced FLU removal in the PMS/RM/HA system possibly resulted from the reduction of Fe (III) to Fe (II) by HA on the surface of α-Fe2O3 present in RM (Eq. (2) and.

(13) (3)), which activated PMS, generating reactive radicals, i.e., SO4•− and •OH (Eq. (4) and (5)) (Zou et al., 2013).. NH2OH + ≡FeIII → ≡FeII + NH2O• + H+. (2). ≡FeIII + NH2O•→ ≡FeII + NHO + H+. (3). HSO5− + ≡FeII → SO4•− + ≡FeIII + OH−. (4). HSO5− + ≡FeII → SO42− + ≡FeIII + •OH. (5). To investigate the Fe(III) reduction capacity of HA, the production of dissolved Fe(II) before and after HA addition under anaerobic conditions was investigated. Figure S3 illustrates that in the presence of HA, virtually no Fe(II) dissolution occurs. However, XPS analysis demonstrated that peaks corresponding to Si and Ti on RM surface were not significantly changed before and after HA addition (Figs. 3(a) and (c)), whereas main peak corresponding to Al(2p3/2) shifted from 73.9 to 74.5 eV (Fig. 3(d)), indicating the formation of Al hydroxides after the reaction (Epa et al., 2004). These observations possibly resulted from the formation of Fe(II)-Al(III)-layered double hydroxides during the reaction between the reduced Fe(II) and Al2O3 on the RM surface (Elzinga, 2012). The portion of deconvoluted peak of O1s at 532.1 eV (hydroxide species) certainly increased to 35.8% when it compares with RM alone (25.5%) (Fig. 3(b)). In addition, H 2O peak, “vacancy” peak (Ovac), and O2− peak were observed at 533.8, 531.4, and 530 eV, respectively (Liu et al., 2017). To confirm the surface reduction of Fe(III) to Fe(II) by HA, the Fe region of the XPS spectra was also analyzed. The Fe(2p3/2) narrow-scan spectra of raw RM consisted of three main peaks at 709.5 (Fe(II) species), 710.7, and 712.8 eV (Fe(III) species; Fig. 3(e); Sihn et al., 2019). These results indicate that the proportions of the Fe(II) and Fe(III) surfaces were. 10.4 and. 89.6%, respectively. When HA was added to the RM suspension, an increase in the Fe(II).

(14) surface content (26.8%) was observed, indicating that the Fe(III) surface had been reduced to Fe(II). These results imply that HA can reduce Fe(III) to Fe(II) on the RM surface, thus enhancing PMS activation. However, a significant decrease of the Fe(II) content (from 26.8 to 1.3%) was observed when we added PMS to HA/RM system, suggesting that the reduced Fe(II) surface can be oxidized to Fe(III) after PMS activation. ESR analysis was conducted to provide more direct evidence of radical generation in the PMS/RM/HA system (Fig. 4). We have prepared two different systems using PMS/RM/HA and H2O2/RM/HA at pH 7. No signal was detected in RM/ H2O2/HA system, indicating that H2O2 could not be activated in RM/HA suspension at pH 7. In comparison, apparent signals of DMPO-OH adduct with the intensity ratio of 1:2:2:1 and DMPO-SO4•− adduct were observed in RM/PMS/HA system. The predominant reactive species was further investigated using t-BuOH and i-PrOH because of their different selectivity (i.e., secondorder rate constants) with OH and SO4 . i-PrOH reacts with both OH and SO4 , with the second-order reaction constant. = 1.9×109 M1 s1 and. s1, while t-BuOH can be considered to be more selective toward •OH ( M-1 s-1) than SO4•- (. = 4-7.42 107 M1 = 6.0 x 108. = 4-8.4 x 105 M-1 s-1) (Anbar and Neta, 1967; Buxton et al.,. 1988; Neta et al., 1988). The percentages of •OH and SO4•− that reacted with FLU and the scavenger could be estimated using the bimolecular rate constants and concentrations of each species (Kamagate et al., 2020). According to our calculations (Table S3), t-BuOH (5 mM) is selective for •OH since it will react with more than 98% of •OH. i-PrOH was used at very high concentration (20 mM) to ensure effective and total scavenging of both radical species. Addition of 5 mM of t-BuOH resulted in a slight decrease in FLU removal (Fig. S4), suggesting that hydroxyl radical is not the main reactive species for FLU removal. The slight inhibition observed could be attributed to the reactivity of t-BuOH with SO4•− under our.

(15) experimental conditions (Table S3). The addition of i-PrOH (20 mM) almost completely inhibited FLU degradation, confirming the predominance of the sulfate radical (Fig. S4). In this study, the impact of light on the PMS/RM/HA system was also investigated by performing additional experiments under UV-A (λmax = 365 nm; intensity: 3.4 mW cm-2) and visible light (λ > 400 nm; intensity: 0.5 mW cm-2) irradiations (Fig. S5(a)). In comparison with the results obtained under dark conditions (kapp = 0.577 h-1, Table S2), UV-A irradiation resulted in a faster FLU removal rate (kapp = 1.396 h-1), whereas visible light irradiation had no effect on the removal rate (kapp = 0.495 h-1). This can be explained taking into consideration the inability of visible light to reduce Fe (III) to Fe (II) or activate PMS (Qu et al., 2019; Wang et al., 2017). Despite the enhanced FLU degradation efficiency under UV-A irradiation due probably to UV-A assist PMS activation (Khan et al., 2014; Yang et al., 2010), further experiments using the PMS/RM/HA system were conducted in the dark.. 3.3. Effects of experimental factors on FLU removal The effects of the concentrations of PMS, HA, and RM on FLU removal using the PMS/RM/HA system were investigated under a range of PMS (0.5–1.5 mM), HA (0.05– 2 mM), and RM (0.05–0.2 g L-1) concentrations. The results depicted in Fig. S6(a) and Fig. 5(a) indicate that an increase in PMS concentration results in a higher and faster FLU removal rate (kapp = 0.192 and 1.923 h-1 at 0.5 and 1.5 mM, respectively (Table S2). This can be attributed to the increased production of SO4•− and •OH, as shown in Eq. (4) and (5). However, higher HA concentrations (> 0.05 mM) exhibit a negative effect on the FLU removal rate (kapp = 0.577 and 0.041 h-1 at 0.05 mM and 2 mM, respectively (Fig. S6(b), Fig. 5(b), and Table S2)). These observations support the scavenging role of HA according to Eq. (6) and (7) (Zou et al., 2013)..

(16) NH2OH + SO4•− → SO42− + nitrogenous products. (k < 1.5 × 107 M-1S-1). (6). NH2OH + •OH → OH− + nitrogenous products. (k < 5.0 × 108 M-1S-1). (7). Indeed, nitrogenous products ( 0.5 mg L-1 of NO3 - and NO2-), which possibly resulted from HA degradation as shown in Eq. (6) and (7), were observed at the end of the 6 h reaction. Additionally, increasing the RM dosage, particularly during the early stage of the reaction (Fig. S6(c) and Fig. 5(c)), enhanced the FLU removal (kapp = 0.577 and 0.993 h-1 at 0.05 mg L-1 and 0.2 mg L-1, respectively (Table S2). The effect of HA on the Fe(III)/Fe(II) cycle on the RM surface was also evaluated using two other reductants, i.e., dithionite and sulfite (Fig. S6(d) and Fig. 5(d)). In the presence of dithionite, the FLU removal efficiency was. 40%, including initial removal of 23% via. adsorption while almost no further removal was observed in the presence of sulfite. These demonstrated the superior FLU removal performance of the PMS/RM/HA system, owing to the stronger scavenging properties of dithionite and sulfite toward the radicals ( 5.5×109 M1 s1 (Buxton et al., 1988) and. =. = 5.0×108 M1 s1 (Neta and Huie,. 1985), not available for dithionite) than those toward the HA (Eq. (6) and (7)). Figure S5(b) indicates that pH has a considerable effect on FLU sorption and oxidation in the PMS/RM/HA system. A significant decrease in FLU removal is observed when the initial pH was increased from 6 to 9. At pH > pHPZC-RM (i.e., 7.3), there may be a decrease in the electrostatic repulsion between RM and HSO5−, resulting in a decrease in the FLU removal efficiency. Conversely, at a pH of 6, the electrostatic attractions of both the negatively charged FLU− and HSO5− with the positively charged RM surface are favored. These can enhance the initial adsorption of FLU and PMS activation on RM surface, and consequently, the increase in the FLU removal efficiency..

(17) 3.4. Effect of mixture on removal process and oxidation pathways The effect of the co-existence of FQs (i.e., FLU and CIP) on their respective removal was investigated. Figure 6 depicts that in both single (Fig. 6(a)) and binary mixture systems (Fig. 6(b)), 5 μM of CIP was completely removed using PMS/RM/HA. Additionally, kapp values approximately 11.7 and 11.0 times higher than those for FLU removal in both single (Fig. S7(a)) and binary mixture systems (Fig. S7(b)), respectively, were obtained for CIP, owing to the faster reaction between CIP and SO4•− (. = 1.2×109 M-1 s-1 ; Mahdi-Ahmed and. Chiron, 2014). CIP, which contains cyclopropane and piperazine rings, exhibits high reactivity with ROS via (i) the cleavage of the cyclopropane moiety owing to the loss of CH 2 units and (ii) the strong interactions between SO4•− and the piperazine ring (Guo et al., 2017; Jiang et al., 2016). Furthermore, a slight decrease in FLU adsorption was observed in the binary system (Fig. 6(b)), possibly owing to the competition with CIP. To identify CIP and FLU intermediates in single systems, HPLC-ESI-MS technology was employed; the mass spectra obtained are presented in Table S4. Figure S8 depicts the variation in the total ion current (TIC) chromatogram of the detected intermediates during CIP oxidation. Seven major intermediates with molecular ion masses of m/z 332, 348, 334, 316(1), 316(2), 245, and 202 were identified (Fig. S9) using the ESI-positive mode that enable to suggest a plausible degradation pathway (Scheme 1). After 30 min of reaction, the peak of CIP molecule (m/z 332 at RT 10.61) decreased with appearance of peaks at RT 6.39 and 2.32 corresponding to m/z 334 and 348, respectively. This indicates that (1) the cleavage of the piperazine ring (m/z 332. 334) owing to attack by SO4•− and •OH (Li et al., 2020;. Shah et al., 2019; Zhao et al., 2017) and (2) the hydroxylation reaction (m/z 332. 348). owing to the attack of the quinolone ring by •OH (Deng et al., 2017; Li et al., 2020), occurred simultaneously at the initial oxidation of CIP in the PMS/RM/HA system. Then, we observed.

(18) the decrease in peaks for m/z 334 and m/z 348 after 1 h of reaction with appearance of two peaks at RT 4.02 and 5.53 corresponding to m/z 316(1) or 316(2) (Table S4). These can be attributed to (1) the defluorination that resulted in the formation of m/z 316(1) from m/z 334 (Zhao et al., 2017) and (2) piperazine ring opening + decarboxylation that resulted in the formation of m/z 316(2) from m/z 348. Additionally, a peak at 1.06 RT corresponding to m/z 202 was first detected after 1 h of reaction (Fig. S8). After 3 h of reaction, most of the peaks had disappeared, and two distinct peaks at RT 6.57 and RT 1.06, corresponding to m/z 245 and m/z 202, respectively, were observed. After 5 h, when the reaction reached completion, the only peak that remained corresponded to m/z 202. This study is the first to report the intermediate, m/z 202, as a byproduct of CIP oxidation owing to PMS activation. These results indicate that the loss of a secondary amine nitrogen by m/z 316(1) owing to SO4•− attack could lead to the formation of m/z 245 (Ao et al., 2018), from which m/z 202 could be obtained via C-N bond cleavage, decarboxylation, and hydroxylation. Similarly, m/z 202 could also be formed from m/z 316(2) via the further oxidation of the piperazine ring, coupled with C-N bond cleavage. Figure S10 depicts the variation in the TIC chromatogram of the detected intermediates during FLU oxidation. During FLU removal, three intermediates with molecular ion masses of m/z 262, 296, and 252, and two others with molecular ion masses of m/z 242 and 244 were detected (Fig. S11) from the positive (Fig. S10(a)) and negative (Fig. S10(b)) ESI modes, respectively. Based on the identified products, a plausible degradation pathway was proposed (Scheme 2). The strong signal of m/z 262 (RT 17.51), which could be attributed to the FLU molecule, decreased after 1 h of reaction as a new peak at RT 16.5 corresponding to m/z 296 (positive mode) appeared. Additionally, a relatively small signal appeared at RT 16.4 (m/z 242) in the negative mode. These findings indicate that both (i) the hydroxylation of the quinolone ring + the ring opening on the C C double bond (m/z 262 → 296) (Feng et al.,.

(19) 2017a, 2017b, 2016; Qi et al., 2019) and (ii) the defluorination of FLU (m/z 262. 242). (Feng et al., 2015; Rodrigues-Silva et al., 2013) occurred simultaneously during the initial oxidation of FLU by the PMS/RM/HA system. After 2 h of reaction, a new peak at RT 15.76 corresponding to m/z 252 (positive mode), was observed. This possibly resulted from hydroxylation, ring cleavage, and FLU decarboxylation (m/z 262 reaction progressed (2 h. 252). However, as the. 5 h), the intensity of the peak corresponding to m/z 296. decreased, while the intensity of those corresponding to m/z 252 (RT 15.76, positive mode) and m/z 244 (RT 16.55, negative mode) increased. These results indicate that pathway 2 may be the main degradation pathway through which the formation of m/z 252 from m/z 296 via decarboxylation can be induced. Additionally, via hydroxylation, defluorination, and hydrogen abstraction, m/z 252 could be further converted to m/z 244. This study is the first to report the identification of the intermediate, m/z 244, as a byproduct of FLU oxidation owing to PMS activation.. 3.5 Impact of HW components on the removal performance The kinetics of FLU and CIP removal from HW is illustrated in Fig. 7. No significant FLU removal was observed in HW (kapp = 0.014 h-1, Table S2; Fig. 7 (a)). The inhibitory effect in HW can be attributed to the major dissolved components in HW, including chlorides, phosphates, nitrates, sulfates, and dissolved organic matter. It should be noted that these wastewater components exist at very high loading with respect to the concentration of target compounds (e.g. molar ratio [phosphate]/[FLU] = 126; [chloride]/[FLU] = 3540), which may lead to strong competitive and/or radical scavenging effects. To evaluate the effect of each of these ligands on FLU removal efficiency, the FLU removal by the PMS/RM/HA system in UPW was examined in the presence of inorganic ions.

(20) and organic matter, similar to those present in HW (Fig. 8). The results obtained revealed that chlorides significantly improved the FLU removal as evidenced by an approximately 25-fold higher kapp value (kapp = 14.401 h-1, Table S2), relative to that obtained in the absence of chlorides (kapp = 0.577 h-1). Reportedly, Cl- in water can react with PMS, generating HOCl (Eq. 8; Yang et al., 2019). Moreover, chlorine radicals (Cl•) can be formed in water when chlorides in water react with SO4•− (Eq. 9; Lei et al., 2019). Furthermore, Cl- in moderate concentration as in this study and at neutral pH could react with SO4•−, resulting in the production of •OH (Lutze et al., 2015). This indicates that there was an additional possibility for converting SO4•− into •OH. Therefore, the enhancement of the FLU removal in presence of Cl- can be attributed to the formation of reactive Cl species and/or •OH.. Cl- + HSO5− → SO42− + HOCl. (k = 2.06 × 103 M-1S-1). (8). Cl- + SO4•− → SO42− + Cl•. (k = 3.1 × 108 M-1S-1). (9). On the contrary, the presence of phosphates resulted in a strong inhibition of the FLU removal, whereas nitrate (kapp = 0.267 h-1) and sulfate (kapp = 0.245 h-1) exhibited lower FLU removal inhibitory effects (Fig. 8). This strong inhibitory effect of phosphates can be explained by considering its strong adsorption onto metal oxides, thereby reducing the interaction between the oxidant and solid surface. Indeed, we confirmed this strong binding of phosphate by conducting an independent adsorption set, which showed great affinity of phosphate to RM surfaces up to 5 mgP/g at pH 7. At molecular-level, phosphate binds strongly to Fe(III) oxides through monodentate and/or multidentate–mononuclear surface complexes, and a variety of surface complexation models has been proposed to describe the phosphate removal by mineral surfaces (Khare et al., 2005). Furthermore, phosphate can also.

(21) exhibit strong scavenging effect toward SO4•− and •OH (Eq. (10) and (11); (Buxton et al., 1988; Neta et al., 1988).. SO4•− + HPO42− → HPO4•− + SO42− •. OH + HPO42− → HPO4•− + OH−. (k = 1.2 × 106 M-1 s-1). (10). (k = 1.5 × 105 M-1 s-1). (11). Reportedly, humic substances generally act as ROS scavengers given that some of their electron-rich moieties (e.g., phenolics) can react with radical species (Deng et al., 2017; Jiang et al., 2016). Additionally, other quinone moieties can possibly activate PMS to generate ROS, thereby accelerating the degradation of organic pollutants (Deng et al., 2017; Zhou et al., 2015). To investigate the effect of humic acids on FLU removal, LHA was added to the PMS/RM/HA system (Fig. 8). The results obtained demonstrated that LHA caused a slight decrease in FLU removal efficiency, indicating that humic acids have lesser effect on FLU removal using the PMS/RM/HA system. The strong inhibitory effect of phosphates was also observed in HW containing a binary FQ system, i.e., FLU + CIP (Fig. S12). Therefore, it can be concluded that the presence of phosphates in HW is primarily responsible for the degradation of FLU removal efficiency in the PMS/RM/HA system. Nevertheless, CIP degradation was not affected by the presence of these compounds (Fig. 7(b)) because of its fast reaction with SO4•−. To overcome the inhibitory effect of phosphates in HW, the concentration of PMS was increased from 1 to 5 mM (Fig. 9(a)). Thus, there was a gradual increase in FLU removal efficiency as PMS concentration increased up to 5 mM, at which complete FLU removal was observed. A significant increase in the kapp value was observed at a PMS concentration of 3 mM (kapp = 0.177 h-1, Table S2) and 5 mM (kapp = 0.388 h-1), in comparison with that at.

(22) 1 mM (kapp = 0.014 h-1). This can be attributed to a higher production of SO4•− as PMS concentration increases. Regarding TOC removal, a removal efficiency of only 30% was obtained using 3 mM PMS for 24 h (Fig. 9(b)). However, the sequential addition of 3 mM PMS over seven times enhanced TOC removal, up to approximately 95% after the seventh addition (Fig. 9(b)), suggesting that controlling the PMS concentration, as well as its addition mode, could guarantee the successful mineralization of FQs using the PMS/RM/HA system. During the sequential experiments, the amount of metals leached (i.e., Al, Fe, Si, and Ti) was insignificant (significantly low concentrations during the first reaction), as presented in Table S5. Additionally, no remarkable increase in metal leaching was observed over seven reaction cycles (Fig. S13), indicating that RM was fairly stable in the PMS/RM/HA system.. 4. Conclusion In this study, the effective activation of PMS by the RM/HA system for the oxidative degradation of FLU and CIP in HW was demonstrated, and the associated oxidation pathways were proposed. On the RM surface, HA acted as a reducing agent in the Fe III/FeII cycle, and its addition to the PMS/RM suspension significantly enhanced FQ removal. The results also demonstrated that SO4•− possibly played a more crucial role in FQ removal in PMS/RM/HA systems than •OH. Additionally, the high reactivity between SO4•− and CIP, which resulted in the rapid removal of CIP, was observed under all the experimental conditions employed in this study. Furthermore, the effect of competition on FLU removal depended on the nature of the organic and inorganic compounds present in HW. Particularly, phosphates could significantly inhibit FLU removal owing to their strong adsorption onto the RM surface, as well as their scavenging effect toward SO4•−. However, this inhibitory effect.

(23) could be successfully overcome by controlling both the PMS concentration and its addition mode. Globally, approximately 95% of alumina is produced via the Bayer process, which yields 1–2 tons of RM for each ton of alumina that is produced (Bray et al., 2018). Particularly, in 2019,. 132 million tonnes of alumina were produced globally (World Aluminum, 2019);. thus,. 132−264 million tonnes of RM were generated. Considering this massive quantity of. RM generated annually, as well as the environmental problems associated with its high alkalinity, developing a novel method by which it can be properly utilized as a high valueadded material is considerably desired. In this light, the results obtained in this study (i) highlight the potential application of the RM-HA combination as an efficient PMS activator for the mineralization of FQs in HW under dark conditions, and (ii) provide new insights into the development of innovative and cost-effective water technologies that can be employed to overcome the practical barriers associated with the removal of antibiotics from various wastewaters.. Conflicts of interest There are no conflicts to declare.. Acknowledgements We thank the Ministère de l’Enseignement Supérieur et de la Recherche Scientifique of Ivory Coast for a PhD grant (NC). This work was supported by the Campus France (PHC-STAR) program and the National Research Foundation of Korea (project number NRF2017K1A3A1A21013653 and NRF-2019R1C1C1003316)..

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(33) O O NH NH. HN. N HN. N. OH OH. F O. Defluorination. O. or. NH2. or. NH2. O. - C3H5NO. O. H2N. N. OH. Ring cleavage. HN N. OH. N. N. O. OH. N. O. m/z : 245 (+). O. OH. F. F O. O. N. N. O. O. O. O. m/z : 334 (+). m/z : 332 (+). Decarboxylation. O. C-N cleavage. m/z : 316(1) (+) O. NH. HN. Hydroxylation. N. N. OH. N. Decarboxylation OH F. N. N. OH. OH. C-N cleavage. Ring cleavage O. O. m/z : 348 (+). O O. m/z : 316(2) (+). O. m/z : 202 (+). Scheme 1. Proposed reaction pathways for the degradation of CIP in the PMS/RM/HA system..

(34) N. OH. Defluorination O. O. m/z : 242 (-) Hydroxylation. Hydorxylation. Pathway II. N. N. Decarboxylation. OH. N. OH. O. O. OH. O. Hydrogen abstraction. Ring opening. Defluorination. N. O. O. F. F. F O. O. O. O. m/z : 262 (+). O OH. m/z : 252 (+). m/z : 296 (+). Pathway IV Ring cleavage (-CO2). O. m/z : 244 (-). Ring cleavage (- CO2) HO. NH. N. O. OH F. F O. m/z : 252 (+). O. O. m/z : 252 (+). Scheme 2. Proposed reaction pathways for the degradation of FLU in the PMS/RM/HA system..

(35) (a). H : Hematite B : Boehmite A : Anatase C : Calcite Q : Quartz. (b). H H B. H B. A Q. C. B. H. H H B H. (c). H H. H. H. Fig. 1. (a) XRD, (b) TEM and (c) its enlarged images of raw RM..

(36) Fig. 2. Removal kinetics of FLU in different reaction systems. Conditions: [FLU] 0 = 5 µM, [PMS]0 = 1.0 mM, [HA]0 = 0.05 mM, [RM] = 0.05 g L-1, reaction time = 6 h (1 h for adsorption of FLU on RM surface), pH = 7.0 ± 0.1, V = 200 mL. Dotted lines denote the end of the sorption and adding of HA and/or PMS..

(37) Fig. 3. XPS spectra for (a) Si2p, (b)O1s, (c) Ti2p, (d)Al2p, and (e) deconvoluted and fitted Fe2p of RM before and after reaction with HA. For Fe2p, case of HA/PMS/RM was added. Conditions : [HA]0 = 0.5 mM, [RM] = 0.5 g L-1, [PMS]0 = 10 mM, reaction time = 6 h, pH = 7.0, V = 200 mL..

(38) Fig. 4. ESR spectra of the DMPO-OH and DMPO-SO4•− adducts generated in RM/PMS/HA and RM/H2O2/HA systems. Conditions : [RM] = 0.05 g L-1, [HA]0 = 0.05 mM, [PMS]0 = [H2O2]0 = 1 mM, [DMPO] = 100 mM, pH = 7.0, t = 10 min, V = 20 mL..

(39) Fig. 5. Variation of kinetic rate constants at different (a) PMS, (b) HA, and (c) RM concentrations in removal of FLU by PMS/RM/HA system, (d) comparison of different reductants (i.e., HA, sulfite, and dithionite). Conditions: [FLU] 0 = 5 µM reaction time = 6 h (1 h for adsorption of FLU to RM surface), pH = 7.0 ± 0.1, V = 200 mL, for fixed concentrations: [PMS]0 = 1.0 mM, [HA]0 = [Sulfite]0 = [Dithionite]0 = 0.05 mM, [RM] = 0.05 g L-1..

(40) Fig. 6. Removal of FLU and CIP in (a) single and (b) binary systems. Conditions: [FLU] 0 = [CIP]0 = 5 µM, [PMS]0 = 1.0 mM, [HA]0 = 0.05 mM, [RM] = 0.05 g L-1, reaction time = 6 h (1 h for adsorption of FLU and CIP on RM surface), pH = 7.0 ± 0.1, V = 200 mL. Dotted lines denote the end of the sorption and adding of HA and PMS..

(41) Fig. 7. Removal of (a) FLU and (b) CIP in single system and different water matrices. Conditions: [FLU]0 = [CIP]0 = 5 µM, [PMS]0 = 1.0 mM, [HA]0 = 0.05 mM, [RM] = 0.05 g L1. , reaction time = 6 h (1 h for adsorption of FLU to RM surface), pH = 7.0 ± 0.1, V = 200 mL.. Dotted lines denote the end of the sorption and adding of HA and PMS. Abbreviations: UPW = Ultra-pure water, HW = hospital wastewater..

(42) Fig. 8. Effect of (a) inorganic ions and (b) organic ligands on FLU removal. Conditions: [FLU]0 = 5 µM, [PMS]0 = 1.0 mM, [HA]0 = 0.05 mM, [RM] = 0.05 g L-1, reaction time = 6 h (1 h for adsorption of FLU on RM surface), pH = 7.0 ± 0.1, V = 200 mL, [Phosphate]0 = 150 mg L-1, [Nitrate]0 = 10 mg L-1, [Sulfate]0 = 120 mg L-1, [Chloride]0 =250 mg L-1, [LHA]0 = 40 mgC L-1, pH0 = 7.0 ± 0.1, V = 200 mL. Dotted lines denote the end of the sorption and adding of HA and PMS. Abbreviation: LHA = Leonardite Humic Acid..

(43) (a). (b) (b). Fig. 9. (a) Effect of PMS concentration (1, 3 and 5 mM) on FLU removal by PMS/RM/HA system in HW. Conditions: [FLU]0 = 5 µM, [HA]0 = 0.05 mM, [RM] = 0.05 g L-1, reaction time = 6 h (1 h for adsorption of FLU on RM surface), pH 0 = 7.0 ± 0.1, V = 200 mL. Dotted lines denote the end of the sorption and adding of HA and PMS. (b) Impact of sequential addition of PMS on TOC removal (HW/FLU) in PMS/RM/HA system. 3 mM of PMS was added at each time..

(44) Graphical abstract.

(45) Table 1. Inorganic species and physico-chemical characteristics of hospital wastewater hospital wastewater. pH. 6.8 ± 0.2. Turbidity (NTU). 196 ± 5. Conductivity (µS cm-1). 1340 ± 5. TOC (mg L-1). 50 ± 10. Suspended solid (mg L-1). 20 ± 2. Chloride (mg L-1). 620 ± 10. Nitrate mg L-1). 7±2. Sulfate (mg L-1). 60 ± 10. Phosphate (mg L-1). 60 ± 10.

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