Degradation of antibiotics by advanced oxidation processes: An overview

Journal Pre-proofs Review Degradation of antibiotics by advanced oxidation processes: An overview Jianlong Wang, Run Zhuan PII: DOI: Reference:

S0048-9697(19)35015-6 https://doi.org/10.1016/j.scitotenv.2019.135023 STOTEN 135023

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Science of the Total Environment

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Please cite this article as: J. Wang, R. Zhuan, Degradation of antibiotics by advanced oxidation processes: An overview, Science of the Total Environment (2019), doi: https://doi.org/10.1016/j.scitotenv.2019.135023

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Degradation of antibiotics by advanced oxidation processes: An overview Jianlong Wanga,b,*, [email protected], Run Zhuana

aLaboratory

of Environmental Technology, INET, Tsinghua University, Beijing

100084, P. R. China bBeijing

Key Laboratory of Radioactive Waste Treatment, Tsinghua University,

Beijing 100084, P.R. China

*Corresponding

author at: Energy Science Building, INET, Tsinghua University, Beijing 100084, China. Graphical abstract

Highlights Antibiotics are ubiquitous due to their extensive production and consumption. AOPs are effective to degrade antibiotics in water and wastewater. The recent advance in antibiotics degradation by AOPs was analyzed and summarized. Fenton, ozonation, photocatalytic, electrochemical and ionizing radiation were introduced. Concluding remarks were given and their future perspectives and challenges were discussed. 1

Abstract Antibiotics are becoming emerging contaminants due to their extensive production and consumption, which have caused hazards to the ecological environment and human health. Various techniques have been studied to remove antibiotics from water and wastewater, including biological, physical and chemical methods. Among them, advanced oxidation processes (AOPs) have received increasing attention due to their fast reaction rate and strong oxidation capability, which are effective for the degradation of antibiotics in aquatic environments. In this review paper, a variety of AOPs, such as Fenton or Fenton-like reaction, ozonation or catalytic ozonation, photocatalytic oxidation, electrochemical oxidation, and ionizing radiation were briefly introduced, including their principles, characteristics, main influencing factors and applications. The current applications of AOPs for the degradation of antibiotics in water and wastewater were analyzed and summarized, the concluding remarks were given and their future perspectives and challenges were discussed. Key words: Advanced oxidation processes; antibiotics; ionizing radiation; ozonation; photocatalytic oxidation; Fenton-like oxidation 1.

Introduction Antibiotics are chemical compounds which are applied to treat microbial

infectious diseases, they have widely applied for the treatment of human and animal diseases as well as in aquaculture and livestock feeding (Manzetti and Ghisi, 2014). The extensive use of antibiotics, especially the overuse or abuse of antibiotics has attracted public concern. During the production and application of antibiotics, a large amount of antibiotics-containing wastewater are generated and discharged into the 2

environment, causing serious pollution (Focazio et al., 2008). The residual antibiotics are persistent and difficult to degrade by conventional biological treatment methods (Wang and Wang, 2016; Prado et al., 2009; Kummerer et al., 2000). Therefore, antibiotic were frequently detected in various natural environments (Wang et al., 2019), including river water (Huang et al., 2019), groundwater (Szekeres et al., 2018), surface water (Danner et al., 2019), soil (Cerqueira et al., 2019), sediment (Chen and Zhou, 2014) and drinking water (Sanganyado & Gwenzi, 2019). The long-term occurrence of antibiotics in the natural environments may lead to the generation of antibiotic resistant genes (ARGs) and antibiotic resistant bacteria (ARBs), accelerating the spread of antibiotic resistance, causing threat to human health and ecological systems (Kummerer, 2009). Various techniques have been studied for the removal of antibiotics from water and wastewater, including coagulation, membrane separation, adsorption and biodegradation (Wang and Wang, 2019a; 2018b; Zhuang et al., 2020; 2019a; 2019b; Wang and Zhuang, 2019; 2017). However, they have not been widely applied due to their low removal efficiency and high operational cost. By contrast, advance oxidation processes (AOPs) can degrade antibiotics or convert them to small molecule substances, which could alleviate the inhibitive effect of antibiotics on microorganisms, and enhance their biodegradability and the removal rate (Wang and Wang, 2019b; Hernandez et al., 2002). Advanced oxidation processes use strong oxidation agents, such as hydroxyl radical (·OH), ozone (O3), superoxide radical (O2−·) to degrade organic pollutants (Wang and Wang, 2018a; Wang and Bai, 2017; Wang and Xu, 2012; Buxton et al., 1988). According to the different ways used to produce oxidation agents, AOPs can be classified into Fenton oxidation, photocatalytic oxidation, electrochemical oxidation and so on (Fig. 1). In this review, the degradation of antibiotics by various advanced oxidation processes (AOPs), including Fenton or Fenton-like reaction, ozonation or catalytic 3

ozonation, photocatalytic oxidation, electrochemical oxidation, and ionizing radiation were briefly introduced, their principles, characteristics, main influencing factors and applications for the degradation of antibiotics in water and wastewater were analyzed and summarized, the concluding remarks and future challenges were discussed.

2.

Fenton and Fenton-like process The combination of ferrous salt and hydrogen peroxide is called Fenton reagent

(Fenton, 1894). Fenton oxidation methods are widely used in wastewater treatment. As for Fenton oxidation method, Fenton reagent (Fe2+ and H2O2) are added into wastewater, which can react to form hydroxyl radicals (·OH), as equations (1) – (3). Fe2+ + H2O2 → Fe3+ + ·OH + OH–

(1)

·OH + H2O2 → ·HO2 + H2O

(2)

2·OH → H2O2

(3)

These radicals could oxidize or degrade antibiotics. Fenton oxidation method has advantages, such as higher degradation efficiency and easy operation. Various operating parameters, including pH value, temperature, H2O2 concentration and Fe2+ concentration, all have influence on the treatment efficiency. However, Fenton oxidation has several disadvantages, which is limited to the acidic condition, and large amount of iron-containing sludge will yield which is difficult to dispose. In order to overcome these disadvantages, other catalysts are used to replace Fe2+, which called Fenton-like oxidation process (Wang and Wang, 2018e; Wang et al., 2016b). 2.1 Fenton-like catalysts Although homogeneous Fenton oxidation can effectively degrade organic pollutants, there are some problems in practical application. Firstly, the utilization rate of H2O2 is low, causing low decomposition rate of pollutants. Secondly, homogeneous Fenton require pH at around 3, which is lower than pH of practical wastewater. Adjusting pH value will increase the operational cost. Finally, adding ferrous salt will cause the production of iron-containing sludge, resulting in secondary pollution. Heterogeneous Fenton or Fenton-like process can be performed at a wide range of 4

pH, the catalyst can be utilized circularly, which can avoid the production of iron sludge (He et al., 2016; Nidheesh, 2015; Soon and Hameed, 2011). Heterogeneous Fenton catalysts mainly include: (1) iron minerals, such as magnetite (Xu and Wang, 2012), goethite (Wang et al., 2015), ferrite (Liu et al., 2012), ferrihydrite (Barreiro et al., 2007), schorl (Xu et al., 2013); (2) zero-valent iron (ZVI) (Xu and Wang, 2011; Zhou et al., 2008); (3) other single metal and metallic oxide, such as MnO2 (Saputra et al., 2013), TiO2 (Zhang et al., 2016a), Pd (Yuan et al., 2011); (4) iron- and iron oxide-loaded materials, commonly used supporters include activated carbon (Sekaran et al., 2011), alumina (Ghosh et al., 2012), clay (Djeffal et al., 2014), silica (Martinez et al., 2007), zeolite (Fukuchi et al., 2014), biosorbents (Daud and Hameed, 2010); (5) metal-organic frameworks (MOFs), which are crystalline functional material composed of transition metal ions and organic ligands (Tang and Wang, 2018b; Etaiw and El-bendary, 2012; Lee et al., 2009). These heterogeneous catalysts have been reported for the degradation of antibiotics, for instance, metacycline (Qi et al., 2019), lincomycin (Ouyang et al., 2019), enrofloxacin (Hou et al., 2019), tetracycline (Zhang et al., 2019), oxytetracycline (Pan et al., 2019), sulfamethazine (Tang and Wang, 2018a). 2.2 Catalyst dosage Catalyst dosage is important in Fenton and Fenton-like oxidation process, which has crucial influence on the degradation of organic pollutants. The overdose of catalyst may scavenge hydroxyl radicals (·OH) and inhibit the degradation of pollutants. Moreover, excessive catalyst dosage will raise costs and limit practical application (Wang et al., 2016b). Qi et al. (2019) observed the influence of catalyst dosage on metacycline degradation using CuCo2O4 nano-catalyst for Fenton reaction. The removal efficiency increased from 38.4% to 89.1% when the dosage increased from 5.0 to 12.0 mg. When further increasing to 15.0 mg, the removal rate only slightly increased to 92.5%. Ouyang et al. (2019) applied iron-based catalysts (GFe0.5) for Fenton-like oxidation of lincomycin. When adding 0.01 g/L GFe0.5, the removal of lincomycin reached 93.85% after 90 min. When the dosages of catalyst was increased to 0.05 and 0.1 g/L, lincomycin was completely removed within 10 min. Nasseh et al. (2019) 5

synthesized FeNi3/SiO2 magnetic nano-catalyst and applied to degrade metronidazole by heterogeneous Fenton-like process. The degradation efficiency increased from 40.96% to 84.29% when catalyst dosage increased from 0.005 to 0.1 g/L, because large amount of active sites were provided, which caused the elevation of hydroxyl radical through the decomposition of hydrogen peroxide, and promoted the degradation of organic pollutants. 2.3 H2O2 concentration H2O2 plays an important role in Fenton oxidation, as the dominant source of hydroxyl radicals (·OH). Insufficient H2O2 dosage will cause the lack of hydroxyl radicals (·OH) and reduce degradation efficiency. By contrast, excessive H2O2 dosage is not suitable for the degradation of pollutants (Wang et al., 2016b). The required theoretical H2O2 dosage could be calculated according to the following equation (4):

(

1

5

)

𝐶𝑎𝐻𝑏𝑁𝑐𝑂𝑑 + 2𝑎 + 2𝑏 + 2𝑐 ― 𝑑 𝐻2𝑂2→𝑎𝐶𝑂2 + (2𝑎 + 𝑏 + 2𝑐 ― 𝑑)𝐻2𝑂 + 𝑐𝐻𝑁𝑂2 (4)

Theoretically, one mole of 𝐶𝑎𝐻𝑏𝑁𝑐𝑂𝑑 requires

(2𝑎 +

1

2𝑏

5

)

+ 2𝑐 ― 𝑑 moles H2O2.

Usually, the actual added H2O2 concentration should be higher than the calculated value according to the chemical equation, which can be examined through the preliminary experiments. Nasseh et al. (2019) synthesized magnetic nano FeNi3/SiO2 composite and used it as heterogeneous Fenton-like catalyst for the oxidation of metronidazole. They found that the degradation efficiency of metronidazole firstly increased with increase of H2O2 dosage from 50 mg/L to 150 mg/L, then it decreased when H2O2 dosage reached 200 mg/L. Qi et al. (2019) evaluated the degradation of metacycline using CuCo2O4 as catalyst. The results showed that the removal of metacycline was 43.6%, 54.3% and 95.1%, respectively when H2O2 dosage was 100, 300 and 500 μL. 2.4 pH value In Fenton and Fenton-like processes, pH value is an important parameter for effective treatment (Wang et al., 2016b). In the traditional homogeneous Fenton processes, the suitable pH value is about 3.0, while in the Fenton-like processes, the optimal pH depends on the reaction system, especially the reaction mechanisms which 6

rely on the catalyst performance (Wang et al., 2016b). Elmolla and Chaudhuri (2009) evaluated the Fenton oxidation of the antibiotics, including amoxicillin, ampicillin and cloxacillin. After 60 min reaction time, COD of antibiotics wastewater degraded 49.0%, 57.7%, 81.5%, 76.9% and 75.6% at pH 2.0, 2.5, 3.0, 3.5 and 4.0, respectively. While DOC degradation percent was 33.9, 43.5, 54.3, 50 and 48.4 at pH 2.0, 2.5, 3.0, 3.5 and 4.0, respectively. The best decomposition of amoxicillin, ampicillin and cloxacillin wastewater achieved at pH 3.0. The decrease of degradation rate at pH over 3.0 may be due to the decrease in dissolved iron. Wan and Wang (2016) studied the influence of pH on the degradation of sulfamethazine using Ce0/Fe0-RGO composites as Fenton-like catalyst. The removal efficiency of sulfamethazine decreased as pH increased from 6.0 to 8.3. The change of pH value had influence on the adsorption of sulfamethazine on the catalyst surface. When pH was over 7.42, which is the pKa2 of sulfamethazine, negative charged catalyst would repel anionic form sulfamethazine, decreasing the adsorption and inhibiting the oxidation reaction. Zhang et al. (2019) investigated the degradation of tetracycline using zero-valent iron and Fe0/CeO2 as Fenton oxidation catalyst. The results showed that the degradation efficiency of tetracycline decreased from 93% to about 50% when pH increased from 3.0 to 5.8 and nZVI was used as catalyst. The removal efficiency of tetracycline was over 93% when pH ranged from 3.0 to 5.8 and Fe0/CeO2 was used, exhibiting high reactivity at a wide range of pH values. 2.5 Antibiotics removal by Fenton and Fenton-like oxidation The degradation of antibiotics by Fenton and Fenton-like oxidation were summarized in Table 1.

3.

Ozonation or catalytic ozonation Ozonation or catalytic ozonation is an environmentally-friendly technology for

wastewater treatment (Wang and Bai, 2017)．Ozone with 2.07 V oxidation potential can oxidize a variety of refractory organic pollutants. Ozone molecule can degrade 7

organic pollutants directly. Moreover, ozone can react with water with the help of catalyst to form hydroxyl radicals (·OH), which has stronger oxidation capability, according to equations (5) - (9) (Yargeau and Leclair, 2008). O3 + H2O → 2·OH + O2

k=1.1 × 10-4 L/(mol·s)

(5 )

O3 + OH- → O2-· + HO2·

k=70 L/(mol·s)

O3 + HO2· → 2O2 + ·OH

k=1.6 × 109 L/(mol·s)

(6) ( 7)

O3 + ·OH → O2 + HO2·

(8)

2HO2· → O2 + H2O2

(9)

Therefore catalytic ozonation process can be used to enhance the degradation efficiency of organic pollutants, including homogeneous and heterogeneous catalytic ozonation. In homogeneous catalytic ozonation process, liquid catalysts, especially transition metal ions are used, such as Fe2+, Mn2+, Ni2+, Co2+, Cd2+, Cu2+, Ag+, Cr3+, Zn2+ in reaction solution. These catalysts can excite ozone to generate hydroxyl radicals (·OH) and improve degradation efficiency. In heterogeneous catalytic ozonation process, solid catalysts such as metal oxide, activated carbon, porous materials and their composite materials are added into reaction solution (Kasprzyk-Hordern et al., 2003). 3.1 Ozone concentration The ozone concentration has important influence on the degradation of antibiotics. The mass transfer rate and the volumetric mass transfer coefficient of ozone increases with increase of ozone concentration. More ozone can be absorbed and react with antibiotic molecules, finally improving the decomposition of antibiotics (Zhao et al., 2006; Kornmuller and Wiesmann, 2003). Oh et al. (2016) studied the influence of ozone dosage on degradation of antibitics, they found that tetracycline was degraded more quickly at 7 ppm ozone exposure than at 3 ppm. Iakovides et al. (2019) found that the elimination of antibiotics increased when the ozone dosage increased, including ampicillin, azithromycin, clarithromycin, 8

erythromycin, ofloxacin, sulfamethoxazole, tetracycline, trimethoprim. Paucar et al. (2019) studied the degradation of ciprofloxacin, levofloxacin, clarithromycin and nalidixic acid, they found that antibiotics degradation enhanced when the initial ozone concentration increased. Hollender et al. (2009) explored the effect of ozone dosage on the elimination of various micro-pollutants. Overall, the removal efficiency of selected micro-pollutants increased with increase of ozone dosage. De Witte et al. (2009) investigated the ozonation of ciprofloxacin, and found that the pseudo first-order constants increased with the increase of ozone inlet concentration. 3.2 pH value Generally, ozone can degrade organic pollutants through direct oxidation by ozone molecule in acidic condition. In alkaline condition, organic pollutants are oxidized by both ozone molecule and hydroxyl radicals (·OH) (Ikehata et al., 2006; Yargeau and Leclair, 2008). Thus, the degradation of antibiotics by ozonation depends on the solution pH values. Feng et al. (2016a) found that the degradation of flumequine was faster at higher pH values. The reaction rate constant increased from 0.3772 min−1 to 2.5219 min−1 when pH values increased from 3.0 to 11.0. In alkaline condition, more O3 was transformed to ·OH, and indirect oxidation of ·OH could be more beneficial in decomposition of flumequine than direct oxidation of O3. Moreover, the species of flumequine under different pH also influenced the result. Wang et al. (2012) explored the chloramphenicol (CAP) degradation by ozone in aqueous solution at various initial pH values. The removal efficiency of chloramphenicol (CAP) was 41.4 ± 1.0% and 65.3 ± 3.0%, respectively at initial pH of 2.0 and 8.0. This result may be attributed to more free radicals generated in alkaline conditions. However, the removal rate decreased at initial pH of 10.0. Oncu and Balcioglu (2013) investigated the influence of pH on the ozonation of ciprofloxacin (CIP) and oxytetracycline (OTC). Higher degradation of ciprofloxacin (CIP) and oxytetracycline (OTC) was achieved at higher pH. Jung et al. (2012) examined the effect of pH on the degradation efficiency of ampicillin, the biodegradability and toxicity after ozonation. The second-order rate 9

constant and COD removal rate increased with increase of pH when pH was in the range of 5–9. A higher biodegradability and acute toxicity was observed at the highest pH (pH 9). On the one hand, an increase of ozone decomposition to generate •OH occurred in alkaline condition. On the other hand, non-protonated organic amine species (-NH2) was more reactive toward ozone molecules than the mono-protonated form (-NH3) (Hoigne & Bader, 1983). At pH 9, non-protonated amine (-NH2) was the dominant group of ampicillin, which could be attacked by ozone more easily. 3.3 Mineralization of pollutants Usually, the ozonation process could not totally mineralize the antibiotics. On the one hand, carbonate (CO32-) and bicarbonate (HCO3-) formed during antibiotic decomposition process are hydroxyl radical scavengers, which can inhibit the removal of antibiotics. On the other hand, solution pH decreased with the ozonation reaction proceeding, which is adverse for the generation of hydroxyl radicals (·OH). Uslu and Balcioglu (2008) observed that mineralization rate of oxytetracycline reached 20% after 30 min ozonation at pH 8.5. Feng et al. (2016a) found that 39.45% of TOC was removed after the ozonation of flumequine aqueous solution. Kuang et al. (2013) observed complete trimethoprim degradation after ozonation, while no mineralization was determined. Goncalves et al. (2012) found that TOC removal efficiency was 33.5% after 180 min ozonation of sulfamethoxazole solution. 3.4 Biodegradability improvement of pollutants The BOD5/COD ratio is usually used for characterizing the biodegradability of a pollutant or wastewater. The biodegradability of antibiotics wastewater can be improved by ozonation due to the generation of low molecule weight and biodegradable intermediate products. Balcioglu and Otker (2003) observed that biodegradability of wastewater after ozonation increased. More low-molecular weight intermediate products that are more amenable to biodegradation generated after ozonation (Stockinger et al., 1995). Jung et al. (2012) found that the BOD5/COD ratio at 9 increased constantly from 0 to 0.41 after 120 min of ozonation, enhancing the biodegradability and biological 10

treatability of ampicillin-containing wastewater. Dantas et al. (2008) reported that the biodegradability increased from 0 to 0.3 during sulfamethoxazole ozonation, indicating that the antibiotic was conversed to biodegradable intermediate product. Uslu and Balcioglu (2008) observed that the BOD5/COD ratio of synthetic oxytetracycline wastewater increased from 0.05 to 0.3 due to the formation of biodegradable intermediate products during ozonation process. 3.5 Antibiotics removal by ozone oxidation The degradation of various antibiotics by ozonation was summarized in Table 2. 4.

Photocatalytic oxidation Photocatalytic oxidation has been extensively studied for the degradation of

organic pollutants. Semi-conductor materials, such as TiO2, ZnS, WO3 and SnO2 are used as photo-catalyst. When photo-catalysts absorb energy, they excites to generate electrons (e-) with high reducing ability and holes (h+) with high oxidizing ability. O2 can be reduced to form superoxide radical (·O2−) by the excited electrons (e-). While holes (h+) migrate to the surface of the photo-catalysts, H2O will be oxidized to generate hydroxyl radical (·OH). Then, the organic pollutants could be decomposed by superoxide radical (·O2−) or hydroxyl radical (·OH). TiO2 is the most commonly used catalyst for its high catalytic efficiency, stability and no secondary pollution. Its mechanisms of photocatalytic oxidation are as following equations (10) – (16) (Liu et al., 2018b; Saadati et al., 2016). TiO2 + hγ → TiO2 + h+ + e-

(10)

h+ + OH- → ·OH

(11)

h+ + H2O + O2 → ·OH + H+ + ·O2−

(12)

O2 + e- → O2-

(13)

O2- + H+ → HO2-

(14)

2HO2- → O2 + H2O2

(15)

H2O2 + ·O2− → ·OH + OH- + O2

(16)

11

4.1 Photocatalytic materials Photocatalytic reaction means a photochemical reaction and redox process occurring between a photo-catalysts and its surface substrates such as H2O2, O2 and target pollutants under light irradiation. Photo-catalysts are very important. To enhance the efficiency of photo-degradation process, various photo-catalysts, including metal oxides (TiO2, ZnO), metal sulfides (such as CdS); precious metal semiconductors (Ag3O4, BiOBr, BiOCl, BiVO4, GdVO4, SmVO4); non-metallic semiconductors (g-C3N4) have been tested in photochemical oxidation (Li et al., 2019; Shandilya et al., 2019; 2018; Sivakumar et al., 2018; Tilley, 2019; Malathi et al., 2018; Qi et al., 2017; Priya et al., 2016; Wen et al., 2015; Zangeneh et al., 2015). Owing to high photocatalytic activity, non-toxicity and high photo-stability, titanium dioxide (TiO2) photo-catalysts have been extensively applied in environment remediation, especially for the degradation of toxic organic pollutants, such as antibiotics (Wen et al., 2015; Kanakaraju et al., 2014), such as levofloxacin (Kansal et al., 2014), oxytetracycline (Espindola et al., 2019), tetracycline (Lyu et al., 2019), ciprofloxacin (Zeng et al., 2019), sulfaquinoxaline (Sandikly et al., 2019). BiVO4 showed excellent photocatalytic activity in degradation of pollutants because it has low band gap, good dispersibility, non-toxicity, resistance to corrosion (Malathi et al., 2018). It has been applied for decomposition of antibiotics, such as ciprofloxacin (Chen et al., 2018; Yan et al., 2013), oxytetracycline (Ye et al., 2019), norfloxacin (Du et al., 2019), penicillin (Liu et al., 2019a), tetracycline (Wang et al., 2019a). Zinc oxide (ZnO) has also been explored for degradation of antibiotics, such as ciprofloxacin (Sarkhosh et al., 2019), norfloxacin (Mamba et al., 2018), sulfamethoxazole (Mirzaei et al., 2018), cefixime trihydrate (Shooshtari and Ghazi, 2017), tetracycline hydrochloride (Ji et al., 2018) because of its low cost, high redox potential, non-toxicity, and environmentally-friendly properties. Various new carbon materials, such as graphene (Sun and Chang, 2014), carbon nanotubes (Yan et al., 2015), fullerene (Ge et al., 2019), carbon quantum dots (Yi et al., 2018) has been investigated for photocatalytic degradation of organic pollutant. They 12

exhibited high catalytic activity in decomposition of antibiotics, such as sulfamethazine (Liu et al., 2019b), oxytetracycline (Wang et al., 2019b), tetracycline (Yuan et al., 2019), levofloxacin (Kaur et al., 2019), and cefixime (Sheydaei et al., 2018). 4.2 pH value Solution pH is important for photocatalytic oxidation. When pH is lower or higher than the potential of zero charge of catalysts, catalyst surface has different charges. Similarly, when pH is lower or higher than the pKa of substrates, substrates show different charged form (Mehrjouei et al., 2015). Dimitrakopoulou et al. (2012) applied UV-A/TiO2 photo-catalyst to degrade amoxicillin and found that amoxicillin degradation showed no apparent change at pH 5.0 and 7.5. The mineralization of amoxicillin decreased from 95% to 75% when pH increased from 5 to 7.5. This may be associated with the ionization states of both the catalyst and pollutants. Leon et al. (2017) investigated the degradation of cefotaxime under sunlight radiation usingTiO2 and ZnO in aqueous solutions. During the cefotaxime degradation using TiO2 as catalyst, cefotaxime removal rate increased when pH increased from 4 to 6.2, then it declined when pH further increased from 6.2 to 7.6. Similar results were also observed during the cefotaxime degradation using ZnO as catalyst. Palominos et al. (2008) investigated the photocatalytic oxidation of flumequine (FQ) using TiO2 as catalyst. The maximum degradation rate of flumequine was observed at medium pH value. At lower pH, catalyst and flumequine (pKa 6.35) are positively charged, inhibiting the affinity between them, while at pH higher than pKa, both catalyst and flumequine are negatively charged. 4.3 Catalysts dosage Photo-catalyst dosage is important in photocatalytic oxidation process. Increasing photo-catalyst loads would increase the reactive sites and then improve the oxidation and mineralization efficiencies of pollutants. However, excessively added catalysts would block the penetration of the photons and cause the loss of light energy through shielding, reflection and scattering of light by solid particles (Gong and Chu, 2015; Nezamzadeh-Ejhieh and Shams-Ghahfarokhi, 2013). 13

Ahmadi et al. (2017) observed the effect of MWCNT/TiO2 dosage on photocatalytic degradation of tetracycline (TC). It was found that tetracycline removal increased when the catalyst dosage increased from 0.1 g/L to 0.2 g/L. While it further increased to 0.4 g/L, tetracycline degradation efficiency did not further increase. Zhang et al. (2018) found that TiO2 dosage strongly influenced the photocatalytic degradation of chloramphenicol (CAP). When TiO2 dosage increased from 0 to 1 g/L, the kinetic rate constant increased from 0.00534 to and 0.03160 min−1. The excess TiO2 may result in light screening effects and decrease light intensity. Lofrano et al. (2014) observed that the photocatalytic degradation rate constant of vancomycin B was 0.013 and 0.036 min−1, respectively at TiO2 dosage of 0.1 g/L and 0.2 g/L. 4.4 Catalysts stability The photo-catalytic efficiency is related to the stability of catalysts. In practical application, the stability of catalysts is an important factor considering economic costs. Zhu et al. (2018) investigated the stability of CdS/Fe3O4/g-C3N4 during the cyclic photocatalytic degradation of tetracycline, and found that the catalytic activity kept almost unchanged for five cycle experiments. Wu et al. (2016) evaluated the stability of photo-catalyst CdS/SrTiO3 heterojunction during five consecutive cycles of photocatalytic degradation of ciprofloxacin, and found that it was stable and photocorrosion resistant. Xue et al. (2015) studied the degradation of tetracycline using Au/Pt/g-C3N4 as photo-catalyst, they found that the photocatalytic efficiency declined 8.7% after four cycles, indicating that the photo-catalyst was stable. Liu et al. (2018b) conducted the photo-catalyst recycling experiments for tetracycline degradation to evaluate the reliability of photo-catalyst BGC1. 4.5 Mineralization of antibiotics The mineralization efficiency is usually lower than degradation efficiency because there are some transient organic intermediates formed during the photocatalytic process. Liu et al. (2016) found that the TOC removal was very slow, it was only 9.5% for oxytetracycline degradation under UV irradiation for 10 h. Ahmadi et al. (2017) found that TOC removal of tetracycline reached to 83% after UV irradiation for 300 min with 14

the addition of 0.2 g/L MWCNT/TiO2. Palominos et al. (2008) observed that the mineralization ratio of flumequine reached 74% after UV irradiation for 15 min, it remained almost unchanged after 60 min even a longer irradiation period. After longer irradiation times, some intermediates such as aromatic ring remained intact. 4.6 Antibiotics removal by photocatalytic oxidation Table 3 summarized the antibiotics removal by photocatalytic oxidation using various catalyst materials.

5 Electrochemical oxidation Electrochemical oxidation is a process in which organic substance are oxidized and converted or decomposed into non-toxic and harmless substance under the action of an electric current (Comninellis, 1994; Martinez-Huitle and Brillas, 2009). The electrochemical oxidation technology includes direct oxidation and indirect oxidation, they generally exist simultaneously (Martinez-Huitle and Ferro, 2006; Simond et al., 1997; Comninellis, 1994). During direct oxidation process, organic matters in water could directly react with anode and lose electrons to form small molecular compounds (Comninellis, 1994; Kirk et al., 1985). As for indirect oxidation process, anions in the water react with anode to produce intermediate products with strong oxidizing ability, and these intermediate products further oxidize and decompose organic substances (Do and Yeh, 1996; Chiang et al., 1995). This process is related to electrolytes. Different electrolytes will produce different strong oxidative products and result in different degradation efficiency (Feng et al., 2016b). Hydroxyl radical (·OH) is a kind of intermediate oxidant generated by indirect electrochemical oxidation, which is adsorbed onto the anode surface. Organic matters also could be oxidized by hydroxyl radicals (·OH) into small molecular compounds and carbon dioxide (Garcia-Segura et al., 2018). 5.1 Electrode materials The earliest anode applied in electrochemical oxidation is a metal electrode, which is a bare electrode without oxide film on its surface. Such anodes are highly conductive, 15

however, they are prone to dissolution during electrolysis process, resulting in anode loss and solution contamination by new impurities. To avoid its disadvantage and improve oxidation efficiency, plenty of new anode materials are studied, including graphite (Liu and Jiang, 2005), glassy carbon (Brimecombe and Limson, 2006), conductive-diamond (Canizares et al., 2006), activated carbon-steel (Canizares et al., 1999), Pt (Rao and Dube, 1996), TiO2 (Zhang et al., 2014), nanostructured TiO2 (Tian et al., 2008), β-PbO2 (Wu and Zhou, 2001), IrO2/Ti (Bonin et al., 2004), Ti/TiO2 -RuO2 -IrO2 (Rajkumar and Palanivelu, 2003), Ti/Pt (Vlyssides et al., 2004),TiO2/Ti/Ta2O5 - IrO2 (Asmussen et al., 2009), Sb dopedSnO2 (Zhao et al., 2009), BiOx - TiO2/Ti (Park et al., 2009). The property of anode is related to the preparation method. The composition ratio, particle size, surface structure, specific surface area and bonding force, all affect the performance of the anode (Feng et al., 2016b). Electrochemical oxidation has been applied for the degradation of antibiotics, such as chlortetracycline (Kitazono et al., 2017), cefazolin (Kitazono et al., 2017), tetracycline (Liu et al., 2015; Miyata et al., 2011), ofloxacin (Jara et al., 2007), lincomycin (Jara et al., 2007), sulfamethoxazole (Eleoterio et al., 2013), trimethoprim (Eleoterio et al., 2013), nitrofurazone (Kong et al., 2015), metronidazole (Kong et al., 2015), ceftriaxone (Li et al., 2018). 5.2 Current density The current density affects the driving force of the electrochemical oxidation reaction, thus it affects the electrochemical oxidation efficiency (Moreira et al., 2017). Kitazono et al. (2017) studied the electrochemical oxidation of chlortetracycline using Ti/PbO2 as anode, they found that the degradation of chlortetracycline followed the pseudo first-order kinetics, and the reaction rate constant increased with increase of applied current density, due to the higher ·OH yield rate under higher current density. Eleoterio et al. (2013) investigated the effect of current density on the removal rate of COD of the wastewater contained antibiotics, such as sulfamethoxazole and trimethoprim. They observed that the COD removal efficiency increased when the current density increased from 10 to 100 mA/cm2. Dirany et al. (2010) studied the effect of current density on the degradation of sulfamethoxazole. Haidar et al. (2013) studied 16

the degradation of sulfachloropyridazine using a boron-doped diamond (BDD) anode, and found that the time required for the complete antibiotic decomposition decreased with the increase of current density. Moreover, the mineralization rate was 76%, 84%, 89%, 93% and 95% when the current density was 100, 200, 300, 350 and 400 mA, respectively. Moreira et al. (2014) observed that trimethoprim was degraded rapidly at high current density, and the kinetic rate constant of trimethoprim degradation increased with increase of current density. 5.3 pH value Solution pH influences the performance of electrochemical oxidation. The improvement or inhibition effect is related to water composition, reaction system (Moreira et al., 2017; Moreira et al., 2014; Almeida et al., 2012). Moreira et al. (2014) investigated the effect of pH on trimethoprim degradation using electrochemical oxidation, they found that at pH 3.0, the presence of HSO4- in solution may scavenge hydroxyl radical (·OH) and reduce the degradation efficiency of trimethoprim; at pH 4.5, the removal rate decreased due to iron precipitation. El-Ghenymy et al. (2013) explored the degradation of sulfamethazine using BDD anode. The mineralization of sulfamethazine reached 90%, which was highest at pH 3.0. DOC was only reduced by 84%, 84%, 80% and 65% at pH of 2.0, 4.0, 5.0 and 6.0, respectively. At pH 2.0, H2O2 is easily reacts with H+ to form peroxonium ion (H2O2+), which makes H2O2 more electrophilic and decreases its reactivity with Fe2+ in Fenton reaction. When pH was over 4.0, the gradual precipitation of Fe3+ reduced the formation of hydroxyl radical (·OH) and inhibited the sulfamethazine removal. Wang et al. (2016c) studied the electrochemical oxidation of ciprofloxacin using SnO2-Sb/Ti electrode. They found that the removal of ciprofloxacin and COD was slightly higher at higher pH. The kinetic rate constant and average current efficiency were maximal at pH 3, higher than that at pH 5, 7, 9 and 11. 5.4 Antibiotics removal by electrochemical oxidation Antibiotics removal by electrochemical oxidation was summarized in Table 4.

17

6 Ionizing radiation The ionizing radiation (including gamma ray and electron beam) is an emerging technology for the degradation of organic pollutants, either through indirect way or direct way (Fig. 2). During water radiolytic process, various active species are formed as equation (17). + H2O → ·OH (2.7) + eaq−(2.6) + ·H (0.55) + H2 (0.45) + H2O2 (0.71) + H3O (2.6)

(17)

The numbers in brackets are chemical yield (G-value), presenting the number of species formed when absorbed 100 eV energy at a pH range of 6.0 - 8.5. Hydroxyl radicals (·OH) can oxidize the organic pollutants, and solvated electrons (eaq−) can reduce the organic pollutants (Wang and Chu, 2016; Wang and Wang, 2007). The degradation of antibiotics by ionizing radiation is influenced by various factors, such as the absorbed dose, initial pH, organic matters and water matrix (Yu et al., 2010a; 2010b; Hu and Wang, 2007).

6.1 Absorbed dose The absorbed dose considerably affects the degradation rate of antibiotics. Generally, the antibiotics degradation increases with increase of absorbed dose (Zhuan and Wang, 2019a; 2019b). The radiation-induced degradation of antibiotics follows the pseudo first-order kinetics model (equations (18) and (19)).

C  C0 e  kD (18) ―ln

𝐶 = 𝑘𝐷 𝐶0

(19)

where C0 and C are pollutant concentration before and after radiation (mg/L); D is absorbed dose (kGy); and k is dose constant (kGy−1). The degradation kinetics of various antibiotics followed to pseudo first-order kinetic model, such as cefaclor (Yu et al., 2008), diclofenac (Zhuan and Wang, 2020a; He et al., 2014), sulfamethazine (Chu et al., 2015) and sulfamethoxazole (Zhuan and Wang, 2020b). 6.2 pH value 18

The pH value has significant influence on the degradation of antibiotics by ionizing radiation (Wang and Wang, 2019c; 2018c). Solution pH can affect the reactive radical composition by equations (20) – (22). eaq− + H+ → ·H

(k = 2.3 × 1010 L/(mol s))

eaq− + ·OH → OH−

(20) (21)

·OH + OH− → ·H2O +O−

(k = 1.3 × 1010 L/(mol s))

(22)

At acidic condition, H+ concentration was higher than OH– concentration, which can combine with eaq− at a rate constant of 2.3×1010 L/(mol s), and inhibit the reaction between eaq− and ·OH. As a consequence, more ·OH would react with antibiotic molecules (Guo et al., 2012). At alkaline condition, OH– concentration was higher than H+ concentration, which can react with ·OH at a rate constant of 1.3×1010 L/(mol s) and form weak oxidative O− and H2O, reducing ·OH concentration and resulting in lower degradation efficiency (Basfar et al., 2005; Sayed et al., 2016a). The acidic condition was more effective than alkaline condition for the degradation of sulphadiazine (Rivas-Ortiz et al., 2017; Guo et al., 2012), ciprofloxacin (Sayed et al., 2016a; Guo et al., 2015b), metronidazole (Santoke et al., 2009), norfloxacin (Sayed et al., 2016b). For the antibiotics with zwitter ion character, solution pH can affect the distribution of their molecular and ionic forms, as well as the surface charge property (De Bel et al., 2009), which will generate attraction or repulsion force between different antibiotic forms, finally affecting the degradation efficiency. Zhuan and Wang (2019a) found that when pH was higher than pKa2 (5.7), sulfamethoxazole was mainly in negative charged forms, which would produce repulsion force, decreasing the reaction rate. 6.3 Inorganic anions, organic matters and matrix Practical waters are complex matrices which usually include anions (such as Cl-, CO32-, HCO3- , NO3−, NO2−) and organic matters (such as humic acid). These compouds may interfere with the radiolytic degradation of antibiotic by reacting with the radical species as equations (23) - (44) (Peñalver et al., 2013; Buxton et al., 1988). 19

ClHO·− + eaq− → Cl− + HO− ClHO·− → Cl

−

(k = 1.0× 1010 L/(mol s))

(k = 6.1× 109 L/(mol s))

+ ·OH

·OH + CO32− → ·CO3−+H2O

(23) (24)

(k = 8.5 × 106 L/(mol s))

(25)

eaq− + CO32− →HCO32−

(k = 6.0 × 10 5 L/(mol s))

(26)

·OH + HCO3−→ ·CO3−+OH−

(k = 3.9 × 10 8 L/(mol s))

(27)

H· + HCO3− → H2 + ·CO3−·

(k = 4.0 × 10 4 L/(mol s))

(28)

eaq− + HCO3− →·CO33−

(k = 3.9 × 10

5

L/(mol s))

(29)

·OH + NO2− →·NO2 +HO−

(k = 6.0× 109 L/(mol s))

(30)

·H + NO2− → NO + HO−

(k = 7.1× 108 L/(mol s))

(31)

eaq− + NO2− → ·NO22−

(k = 3.5× 109 L/(mol s))

H+ + NO3− → HNO3

(k = (4.4-6.0)× 108 L/(mol s))

(k <= 1.0× 107 L/(mol s))

·NO3 + H2O → HNO3 + ·HO eaq− + NO3− → ·NO32−

(33)

(k = (0.88-1.2)× 108 L/(mol s)) (34)

·OH + HNO3 → H2O + ·NO3 ·H + HNO3 → H2 +·NO3

(32)

(k = 3.0× 102 L/(mol s))

(k = 9.7× 109 L/(mol s))

(35) (36) (37)

·NO32− + H+ → NO2 +HO−

(k = 4.5× 1010 L/(mol s))

(38)

·H + NO3− → ·NO2 +HO−

(k = 4.4× 106 L/(mol s))

(39)

·NO2 + ·H → HNO2

(k = 1.0× 1010 L/(mol s)) (k = 1.0× 1010 L/(mol s))

·OH + ·NO3 → HONO3 ·H + ·NO3 → HNO3 Cl− + ·OH → ClHO·−

(k = 1.0× 1010 L/(mol s)) (k = 4.3× 109 L/(mol s))

ClHO ·− + H+ → Cl· + 2H2O

(k = 2.1× 1010 L/(mol s))

(40) (41) (42) (43) (44)

The presence of CO32− reduced the removal rate of sulphadiazine (Guo et al., 2012), ciprofloxacin (Guo et al., 2015c), norfloxacin (Sayed et al., 2016b), amoxicillin (Wang et al., 2017b). The presence of HCO3− had inhibitory effect on the degradation of antibiotics, such as amoxicillin (Wang et al., 2017b), sulfamethoxazole (Zhuan & Wang, 2019b), ofloxacin (Wang et al., 2017b), norfloxacin (Sayed et al., 2016), cefradine (Wang et al., 2017b). The presence of NO3− and NO2− decreased the decomposition of 20

antibiotics, such as ciprofloxacin (Guo et al., 2015c), norfloxacin (Sayed et al., 2016), sulfamethoxazole (Zhuan and Wang, 2019b). Wang and Wang (2018d) studied the radiation-induced degradation of sulfamethoxazole in the presence of various inorganic anions, including chloride, bicarbonate, carbonate, nitrate, sulfate and phosphate. The results showed that inorganic anions had obvious influence on SMX degradation, which was dependent on their initial concentrations, suggesting that the effect of inorganic anions on the radiation-induced degradation of sulfonamides antibiotics should be considered when radiation technology is used for the treatment of industrial wastewater. The existence of humic acid could decrease the degradation efficiency of various types of antibiotics, such as ciprofloxacin (Guo et al., 2015c), amoxicillin (Wang et al., 2017b), sulfamethoxazole (Zhuan and Wang, 2019b), ofloxacin (Wang et al., 2017b), cefradine (Wang et al., 2017b), fluoroquinolone (Tegze et al., 2018).

21

7 Concluding remarks and perspectives Antibiotics are becoming emerging contaminants, which have received increasing attention in recent years because they are ubiquitous in the natural environment. Moreover, antibiotics can transfer and accumulate through food chain. The long-term existence of antibiotics in the environment will cause the generation of antibiotic resistance genes (ARGs) and antibiotic resistant bacteria (ARBs), posing potential threat to ecosystem and human health. Due to high degradation efficiency and rate, advances oxidation processes are promising for degradation of antibiotics in water and wastewater. The advantages and disadvantages of different advanced oxidation processes for antibiotics removal were analyzed, summarized and compared in supporting information (Table S1). At present, research on antibiotics removal by advanced oxidation processes have made some progress, and the future research should be focused on the following aspects. (1) Advanced oxidation processes need to be optimized to improve their adaptability and practicability, such as enhancing the efficiency of the catalysts, and the utilization efficiency of ozone or H2O2. (2) Degradation effect of antibiotics by advanced oxidation processes has been investigated. The generation mechanism of free radicals and the degradation mechanism of pollutants are not yet clear. More attention should be paid to the mechanism study. (3) Advanced oxidation processes can effectively degrade antibiotics in water and wastewater, their potential for the removal of ARGs and ARBs has not been studied, which needs further investigation. (4) Actual wastewater is complicated, usually containing multiple antibiotics and other organic pollutants, as well as inorganic compounds, which may decrease the degradation efficiency of antibiotics compared with single antibiotic in aqueous solution. Thus, more studies are needed to pay attention to the practical wastewater and finally fulfil the industrial application. (5) It is difficult to efficiently treat the complicated antibiotic wastewater by only advanced oxidation processes. Combining AOPs with biological treatment methods could be one way to resolve this problem, especially to enhance the mineralization of pollutants. The integrated treatment methods can reduce the operational costs and improve the processing efficiency. (6) The operational cost is crucial for the practical applications, how to reduce the treatment cost of AOPs is also important, and their cost-effect analysis should be considered in future studies.

Declaration of Competing Interest 22

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.

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41

Fig. 1. Hydroxyl radicals (·OH)-based advanced oxidation processes Fig. 2. Principles of ionizing radiation for the decomposition of organic pollutants Table 1. Antibiotics removal by Fenton and Fenton-like oxidation

Antibiotics

Amoxicillin

Ampicillin Azithromycin Cloxacillin Ciprofloxacin Clarithromycin Chlorpheniramine Doxycycline Lincomycin Metacycline Metronidazole Nofloxacin Ofloxacin Oxytetracycline Sulfamethazine Sulfamethoxazole Sulfadiazine Sulfathiazole Tetracycline

Catalyst(dosage); pH range zero-valent iron (nZVI) (0.2 - 2g/L); pH = 2 - 5 Fe(II) (0.32 - 24.3 mM); pH = 2 - 4 H2O2/Fe2+= 2.0 – 150; pH = 2.0 - 4.0 H2O2/Fe2+ = 1-50; pH = 1 -9 Fe(II) (53 - 87 μM); pH = 2.3 - 5.7 Fe(II) (0.32 - 24.3 mM); pH = 2 - 4 H2O2/Fe += 2.0 – 150; pH = 2.0 - 4.0 H2O2/Fe2+= 1.75 mM; pH = 3 Fe(II) (0.32 - 24.3 mM); pH = 2 - 4 H2O2/Fe 2+= 2.0 – 150; pH = 2.0 - 4.0 Fe3O4 (1.0 - 2.5 g/L); pH = 3 - 11 CNTs/FeS (5- 35 mg); pH = 1- 12 H2O2/Fe2+= 1.75 mM; pH = 3 H2O2/Fe2+= 1.75 mM; pH = 3 nZVI (11.2 - 28 g/L); pH = 2 - 5 SBC@b-FeOOH GFe0.5 (0.01 g/L) CuCo2O4 (0.1 – 0.3 g/L) FeNi3/SiO2 (0.005 - 0.1 g/L); pH = 3 - 11 [Fe(II)] (0.8 - 3 mM) Alg/Fe (0.2 - 1.4 g/L); pH = 3 Alg/CDTA/Fe (0.01 - 0.09g); pH = 3 CQDs/Cu-MIO (0.1 - 0.25 g/L); pH = 3.6 - 10 Fe-Cu@MPSi (0.5 - 1g/L); pH = 3 - 9 Cu@Fe3O4 (0.1 - 1g/L); pH = 3.10 - 9.05 Fe0 (0.3 mM) Fe0 (0.3 mM) CUS-MIL-100(Fe) (0.2 - 1.5 g/L); pH = 3 - 6 Ce0/Fe0-RGO (0.1 - 1 g/L); pH = 6 -8 Zn-Fe-CNTs (0.2 - 1g/L); pH = 1.0 - 3.0 Fe0 (0.3 mM) Fe3O4/Humic acid (0 - 5 g/L); pH = 3.5 - 9 Fe@Bacillus subtilis (0.5 g/L); pH = 4.0 - 6.0 42

Removal efficiency (%) 86.5 80 100 80.92 90.2 80 100 95 80 100 89 91.03 95 95 100 100 95.1 95.32 100 100 100 100 100 100 100 100 99 100 100 100 100

References (Zha et al., 2014) (Elmolla et al., 2010) Elmolla and Chaudhuri (2009) (Guo et al., 2015a) (Rozas et al., 2010) (Elmolla et al., 2010) Elmolla and Chaudhuri (2009) (Mackul'ak et al., 2015) (Elmolla et al., 2010) Elmolla and Chaudhuri (2009) (Hassani et al., 2018) (Ma et al., 2015) (Mackul'ak et al., 2015) (Mackul'ak et al., 2015) (Wang et al., 2016a) (Zhang et al., 2016b) (Ouyang et al., 2019) (Qi et al., 2019) Nasseh et al. (2019) (Santos et al., 2015) (Titouhi and Belgaied, 2016a) (Titouhi and Belgaied, 2016b) (Tian et al., 2017) (Zheng et al., 2017) (Pham et al., 2018) (Pan et al., 2019) (Pan et al., 2019) (Tang and Wang, 2018) Wan and Wang (2016) (Liu et al., 2018c) (Pan et al., 2019) (Niu et al., 2011) (Zheng et al., 2016)

Fe0 (0.3 mM) CFO (0.05 - 0.2 g/L) Fe0/CeO2 (0.01 - 0.2 g/L); pH = 3 - 7

100 84 93

(Pan et al., 2019) (Parmar et al., 2017) (Zhang et al., 2019)

Table 2. Antibiotics removal by ozone oxidation

Antibiotics Amoxicillin Ampicillin Chloramphenicol Clarithromycin

Ciprofloxacin Ceftriaxone Ceftriaxone Clindamycin Doxycycline Erythromycin Fumequine Isoproturon Levofloxacin Nalidixic acid Metronidazole Oxytetracycline Ofloxacin Roxithromycin, Sulphadiazine Sulfamethoxazole

O3 dosage or flow rate [O3]0 = 18 mg/L [O3]0 = 0 - 80 mg/L [O3]0 = 5 mg/min [O3]0 = 14 - 42 mg/L O3 dosage of 57 mg/h [O3]0 = 0.125- 0.75 gO3/gDOC [O3]0 = 14 - 42 mg/L [O3]0 = 0.4-1.16 g/g DOC [O3]0 = 3 mg/L [O3]0 = 660 - 3680 ppm [O3]0 = 0.31 – 0.45 g O3/g TS [O3]0 = 2 mg/L [O3]0 = 0 - 80 mg/L [O3]0 = 2.96 g/h O3 flow rate of 0.5 - 1.0 L/h [O3]0 = 3 mg/L O3 flow rate of 0.5L/min [O3]0 = 3 mg/L [O3]0 = 0.125- 0.75 gO3/gDOC [O3]0 = 5 - 7 mg/L [O3]0 = 0.1 - 0.5 mg/L [O3]0 = 140.5 mg/L [O3]0 = 3 mg/L [O3]0 = 14 - 42 mg/L O3 flow rate of = 100 mg/h [O3]0 = 14 - 42 mg/L [O3]0 = 5 - 7 mg/L [O3]0 = 11.2, 32.7 g/m3 [O3]0 = 0.31 – 0.45 g O3/g TS O3 flow rate of 0.5L/min [O3]0 = 0.125- 0.75 gO3/gDOC [O3]0 = 3 mg/L [O3]0 = 0 - 80 mg/L [O3]0 = 50 g/m3 [O3]0 = 0-1.6 g/L [O3]0 = 14 - 42 mg/L [O3]0 = 0.4-1.16 g/g DOC 43

Removal efficiency (%) 99 70 - 98 100 97 >70 100 15-99 >70 95 98 85.4 70 - 98 >95 >70 86.4 - 93.6 >70 >70 43 - 100 100 100 >70 100 >90 100 43 - 100 100 88 86.4 - 93.6 >70 >70 70 - 98 100 100 100 15-99

References (Marcelino et al., 2017) (Alsager et al., 2018) (Jung et al., 2012) (Paucar et al., 2019) (Wang et al., 2012) (Iakovides et al., 2019) (Paucar et al., 2019) (Hollender et al., 2009) (El-taliawy et al., 2017) (De Witte et al., 2009) (Oncu and Balcioglu, 2013) (Lu et al., 2019) (Alsager et al., 2018) (Balcioglu and Otker, 2003) (Norte et al., 2018) (El-taliawy et al., 2017) (Wang et al., 2018) (El-taliawy et al., 2017) (Iakovides et al., 2019) (Ostman et al., 2019) (Michael-Kordatou et al., 2017) (Feng et al., 2016a) (El-taliawy et al., 2017) (Paucar et al., 2019) (Chen et al., 2017) (Paucar et al., 2019) (Ostman et al., 2019) (Uslu and Balcioglu, 2008) (Oncu and Balcioglu, 2013) (Wang et al., 2018) (Iakovides et al., 2019) (El-taliawy et al., 2017) (Alsager et al., 2018) (Goncalves et al., 2012) (Dantas et al., 2008) (Paucar et al., 2019) (Hollender et al., 2009)

Sulfaquinoxaline Trimethoprim

O3 flow rate: 0.4 mL/min [O3]0 = 3 g/h O3 flow rate: 5.52 ± 0.32 mL/min [O3]0 = 2 mg/L [O3]0 = 3 - 7 mg/L

44

99 >85 >99 70 100

(Yin et al., 2017) (Guo et al., 2015b) (Urbano et al., 2017) (Lu et al., 2019) (Oh et al., 2016)

Table 3. Antibiotics removal by photocatalytic oxidation

Antibiotics Amoxicillin Cefixime Cefixime trihydrate Cefotaxime Chloramphenicol

Ciprofloxacin

Enrofloxacin FLumequine Gatifloxacin Levofloxacin Norfloxacin

Oxytetracycline

Penicillin Sulfamethazine Sulfamethoxazole Sulfaquinoxaline

Tetracycline

Vancomycin

Photo-catalysts UV-A/TiO2 N-TiO2/GO ZnO/α-Fe2O3 TiO2/ZnO TiO2 TiO2/carbon dots RGO-BiVO4 Ag/AgBr/BiVO ZnO CdS/SrTiO3 g-C3N4, Fe3O4/g-C3N4 CdS/SrTiO3 g-C3N4, Fe3O4/g-C3N4 TiO2 g-C3N4, Fe3O4/g-C3N4 TiO2 rGO-CdS BiVO4/WO3 ZnO/CuOx TiO2 MWCNT/BiVO4 TiO2@GO GQDs/g-CNNR CdS/SrTiO3 BiVO4 G/A/TNS UVC lamp (10 W) TiO2 g-C3N4, Fe3O4/g-C3N4 Mesoporous TiO2 Au/Pt/g-C3N4 ZnO@ZnS MWCNT/TiO2 BiVO4/Bi2Ti2O7/Fe3O4 TiO2

Removal efficiency (%) 100 80 99.1 84.2 85 91.1 68.2 91.40 100 93.7 100 93.7 100 100 100 90 82.7 70 >80 100 88.8 100 80 100 93.7 100 96.1 95.76 100 100 100

References (Dimitrakopoulou et al., 2012) (Sheydaei et al., 2018) (Shooshtari and Ghazi, 2017) (Leon et al., 2017) (Zhang et al., 2018) (Zeng et al., 2019) (Yan et al., 2013) (Chen et al., 2018) (Sarkhosh et al., 2019) (Wu et al., 2016) (Zhu et al., 2018) (Wu et al., 2016)

80.9 100 97.14

(Zhu et al., 2018) (Palominos et al., 2008) (Zhu et al., 2018) (Kansal et al., 2014) (Kaur et al., 2019) (Du et al., 2019) (Mamba et al., 2018) (Espindola et al., 2019) (Ye et al., 2019) (Wang et al., 2019b) (Yuan et al., 2019) (Liu et al., 2016) (Wu et al., 2016) (Liu et al., 2019a) (Liu et al., 2019b) (Mirzaei et al., 2018) (Sandikly et al., 2019) (Zhu et al., 2018) (Lyu et al., 2019) (Xue et al., 2015) (Ji et al., 2018) (Ahmadi et al., 2017) (Wang et al., 2019a)

95

(Lofrano et al., 2014)

45

Table 4. Antibiotics removal by electrochemical oxidation

Type of antibiotics Cefazolin Ceftriaxone sodium Chlortetracycline Ciprofloxacin Doxycycline Oxytetracycline Sulfamethazine Sulfachloropyridazine Sulfamethoxazole Tetracycline

Anode material Ti/PbO2 RuO2 -TiO2 /Nano-G Ti/PbO2 Ti/IrO2, Ti/PbO2 SnO2-Sb/Ti Ti/IrO2, Ti/PbO2 Ti/IrO2, Ti/PbO2 Boron-doped diamond (BDD) BDD/carbon Pt/carbon, BDD/Carbon Boron-doped diamond Ti/IrO2, Ti/PbO2 Carbon nanotube

46

Removal efficiency (%) 100 >97.3 100 >99 99.5 >99 >99 100 100 100 >99 96.3

References (Kitazono et al., 2017) (Li et al., 2018) (Kitazono et al., 2017) (Miyata et al., 2011) (Wang et al., 2016c) (Miyata et al., 2011) (Miyata et al., 2011) El-Ghenymy et al. (2013) Haidar et al. (2013) Dirany et al. (2010) (Moreira et al., 2014) (Miyata et al., 2011) (Liu et al., 2015)