Pharmaceutically active compounds in aqueous environment: A status, toxicity and insights of remediation

Environmental Research 176 (2019) 108542

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Review article

Pharmaceutically active compounds in aqueous environment: A status, toxicity and insights of remediation

T

Abhradeep Majumdera, Bramha Guptab, Ashok Kumar Guptac,∗ a

School of Environmental Science and Engineering, Indian Institute of Technology Kharagpur, Kharagpur, 721302, India School of Water Resources, Indian Institute of Technology Kharagpur, Kharagpur, 721302, India c Environmental Engineering Division, Department of Civil Engineering, Indian Institute of Technology Kharagpur, Kharagpur, 721302, India b

ARTICLE INFO

ABSTRACT

Keywords: Ecotoxicological impacts Drinking water equivalent limit Degradation efficiency AOPs Pharmaceuticals

Pharmaceutically active compounds (PhACs) have pernicious effects on all kinds of life forms because of their toxicological effects and are found profoundly in various wastewater treatment plant influents, hospital effluents, and surface waters. The concentrations of different pharmaceuticals were found in alarmingly high concentrations in various parts of the globe, and it was also observed that the concentration of PhACs present in the water could be eventually related to the socio-economic conditions and climate of the region. Drinking water equivalent limit for each PhAC has been calculated and compared with the occurrence data from various continents. Since these compounds are recalcitrant towards conventional treatment methods, while advanced oxidation processes (AOPs) have shown better efficiency in degrading these PhACs. The performance of the AOPs have been evaluated based on percentage removal, time, and electrical energy consumed to degrade different classes of PhACs. Ozone based AOPs were found to be favorable because of their low treatment time, low cost, and high efficiency. However, complete degradation cannot be achieved by these processes, and various transformation products are formed, which may be more toxic than the parent compounds. The various transformation products formed from various PhACs during treatment have been highlighted. Significant stress has been given on the role of various process parameters, water matrix, oxidizing radicals, and the mechanism of degradation. Presence of organic compounds, nitrate, and phosphate usually hinders the degradation process, while chlorine and sulfate showed a positive effect. The role of individual oxidizing radicals, interfering ions, and pH demonstrated dissimilar effects on different groups of PhACs.

1. Introduction The presence of pharmaceutically active compounds (PhACs) in the aqueous ecosystem have been known from the mid-20th century. However, in the last few years, with the advent of new analytical technologies and after meticulous studies about the ecological impacts of PhACs, their presence has become an emerging concern. Presence of PhACs in the aquatic ecosystem pose a serious threat to both aquatic and terrestrial organisms. Although these compounds are present in vestigial levels, prolonged exposure can have a pernicious effect on human health, aquatic life, and plants, etc. Unrestrained use of PhACs

such as antibiotics, analgesics, β-blockers, hormones, stimulants, antiepileptics, etc. over the past few decades has polluted many aquatic ecosystems. At present, approximately 3000 compounds are used as medicines with annual production exceeding hundreds of tons (Sim et al., 2011). The worldwide pharmaceutical revenue has increased from 390.2 billion US$ in 2001 to 1105.2 billion US$ in 2016, indicating an increase in pharmaceutical consumption by 2.8 times over the last 15 years (Statista, 2018). The PhACs present in the aquatic ecosystem originates from sources like domestic sewage, hospital effluent, pharmaceutical manufacturing industries, animal husbandries, etc.

Abbreviations: ADI, Acceptable daily intake; AO, Anodic Oxidation; AOP, Advanced oxidation process; BDD, Boron-doped diamond; DWEL, Drinking water equivalent limit; DWI, Daily water intake; e−, Electrons; eV, Electron volt; EEO, Electrical energy per order; EF, Electro Fenton; GAC, Granular activated carbon; h+, Holes; H2O2, Hydrogen peroxide; HOCl, Hypochlorous radicals; kGy, Kilo Grays; kHz, Kilohertz; Kow, Octanol-water partition coefficient; kW, Kilowatt; mA, Milli Ampere; meV, Milli electron volt; OH, Hydroxyl radical; O2, Superoxide radical; PEF, Photo-electro Fenton; PhACs, Pharmaceutically active compound; pKa, Acid dissociation constant; PNEC, Predicted no effect concentration; PMS, Peroxymonosulphate; PS, Persulphate; PZC, Point of zero charge; SHE, Standard hydrogen electrode; SO42-, Sulfate radicals; W, Watt; WWTP, Wastewater treatment plant ∗ Corresponding author. E-mail addresses: [email protected] (A. Majumder), [email protected] (B. Gupta), [email protected] (A.K. Gupta). https://doi.org/10.1016/j.envres.2019.108542 Received 29 January 2019; Received in revised form 12 June 2019; Accepted 17 June 2019 Available online 21 June 2019 0013-9351/ © 2019 Elsevier Inc. All rights reserved.

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The inability of conventional wastewater treatment plants (WWTPs) to completely remove the PhACs lead to the pollution of receiving water bodies. PhACs have been generally found in wastewater influent and effluent, hospital effluent, surface water, and groundwater, etc. Various concentrations of these pharmaceuticals in water bodies have been reported in all prominent countries from different continents (Tran et al., 2017). The rising concern regarding PhACs in WWTP influent, hospital effluent and surface water has provoked research in the field of quantification of these contaminants and their remediation around the globe (Ebele et al., 2017; Tran et al., 2017; Yang et al., 2017a, 2017b). The lack of proper guidelines regarding permissible limits of these contaminants in water has failed to address the gravity of PhAC contamination. A wide range of techniques including advanced oxidation processes (AOPs), adsorption, membrane bioreactor (MBR), moving bed bioreactor (MBBR), etc. have been employed to remove PhACs from the water. However, biological processes like MBR, MBBR take a lot of time to degrade, and some of the PhACs are toxic to the microorganisms, thereby killing them. Adsorption produces contaminated sludge which if not disposed of properly can again contaminate the environment (Chong et al., 2017; He et al., 2017a; Jiang et al., 2017; Lima et al., 2017; Ncibi et al., 2017; Tang et al., 2017). On the contrary, AOPs are cost-effective and less time consuming with a high potential in degrading PhACs. AOPs include processes like ozonation, photocatalysis, Fenton based processes, anodic oxidation (AO), sulfate activated degradation, sonolysis, microwave, plasma discharge, electron beam, etc., which have drawn enormous attention over the past few years. AOPs work on the principle of generating strong oxidizing radicals in the presence of certain catalytic materials, light, electricity, heat, etc. These radicals which when released in the contaminated water degrades the PhACs and mineralizes them ( Wang and Wang, 2018a). Many research articles have shown the occurrence of these PhACs, but they have been limited to only a certain geographical area. Global coverage and the varying concentrations of the individual drugs in their respective geographical area have not been addressed with full transparency. Furthermore, very few researchers have attempted to draw a comparison between the reported concentrations of pharmaceuticals found in water with a proper safe drinking limit (Tran et al., 2017; Yang et al., 2017a, 2017b). The effects of prolonged unmonitored consumption of these pharmaceuticals on the non-target species have not been identified properly. Many research articles have described the removal efficiencies of PhACs by various processes, but the influence of water matrix and other influencing parameters on the removal efficiency of the processes have not been properly addressed (Oturan and Aaron, 2014; Serpone et al., 2017; Wang et al., 2012). A detailed study of the optimum parameters of the treatment system for degrading any particular group of PhACs has also not been properly compiled in any other work. This paper presents a comparative study of the varying concentration of selected PhACs in WWTP influent, hospital effluent and surface water found across the globe and how it compares with the calculated drinking water equivalent level (DWEL) for the respective drug. The removal efficiency, reaction time, and energy required for PhACs degradation by various AOPs have been outlined. The effects of water matrix and other influencing factors on the individual processes in the removal of PhACs along with the optimum conditions of the treatment system for various groups of PhACs have also been presented. Complete degradation of organic compounds is often not achieved. Similarly, the PhACs are not completely mineralized,Instead they are broken down to less harmful or more harmful compounds. The various transformation products formed from different PhACs during AOPs have also been addressed. Furthermore, different AOPs have been compared with respect to treatment time, removal efficiency, and energy consumption.

2. Sources and occurrence The over-utilization and improper management of pharmaceuticals are the primary reasons for the contamination of water. PhACs have been found in WWTP influent, hospital effluent, and surface water. Herein, under WWTP influent, we have included wastewaters from pharmaceutical industries and municipal wastewater. The effluent of WWTPs ends up getting discharged in public sewers, inland surface water, and irrigation land or marine coastal areas. As a result, these effluent get diluted and eventually finds their way into surface water. Herein, under surface waters, we have included lakes, rivers, and any other water bodies where aquatic organisms and microorganisms thrive and are direct or indirect sources of human drinking water. The PhACs have been selected based on their frequency of occurrence in the environment, the threat they possess to the living organisms, the quantity of degradation, and removal studies conducted. The chemical profiles of the selected pharmaceuticals are given in Table 1. Quite a few countries across different continents like Asia (India, China, Singapore, Korea, and Thailand), Africa (Kenya, South Africa), Europe (Greece, Spain, the United Kingdom, Portugal, Italy, Norway, and Germany), and North America (the USA, Canada, and New Mexico) have conducted studies pertaining to occurrence of PhACs in the aquatic ecosystem. The occurrence pattern of the PhACs from these countries was thoroughly studied. 2.1. Sources and persistence of PhACs in the aquatic ecosystem Domestic wastewater containing human excreta and secretions contain high concentrations of PhACs(Nikolaou et al., 2007; Sarmah et al., 2006; Vymazal et al., 2015; Xiong et al., 2016). The non-metabolized part of the antibiotics and analgesics excreted out by humans are found in trace amounts in human urine and feces (Nikolaou et al., 2007; Xiong et al., 2016). Hormones find their way into the wastewater from various sources like human excretion, by-products of synthetic chemicals, plants, and fungi (Hartmann et al., 2014). Women in their menstrual cycle secrete about 10–100 μg of hormones daily, and a pregnant woman secretes about 30 mg of hormones, which eventually escape along with domestic wastewater (Vymazal et al., 2015). βblockers, antibiotics, and analgesics have been found in hospital effluent because of their use on a regular basis (Haro et al., 2017). The sources and pathways of the PhACs in aquatic environments have been depicted in Fig. 1. Most of the PhACs are highly polar and soluble, making them hard to remove. The low octanol/water partition coefficient (log Kow) values of most of the PhACs indicate high solubility. They have very high mobility in the environment and do not get easily hydrolyzed or biodegraded in ambient conditions. Since many of the WWTPs are unable to remove these compounds with good efficiency, these pollutants remain in the effluent (Behera et al., 2011; Daneshkhah et al., 2017). The persistent nature of the PhACs and inability of the conventional WWTPs to completely remove them will increase their concentrations to an alarming level. Also, these PhACs are often termed as “pseudo-persistent” because they have greater environmental persistence than pesticides and dyes. This is a direct result of their source being continually replenishable. PhACs are constantly released into the environment, and the amount in which they are being released is increasing over the years (Ebele et al., 2017). Global annual expenditures on pharmaceuticals have increased from 887 billion USD (United State Dollars) in 2010 to 1135 billion USD in 2017 (Statista, 2018). These PhACs present in the aquatic ecosystem undergo bioaccumulation and are known to bio-accumulate in algae and fishes present in water. As a result, their concentration keeps on increasing as they go higher up the food chain. Carbamazepine was reported to be bio-accumulated in algae (Pseudokirchneriella subcapitata) and crustaceans (Thamnocephalus platyurus) (Ebele et al., 2017). In algae, the high lipid content acts as the entry point for lipophilic PhACs. High concentrations of other PhACs have 2

Diclofenac Codeine Ibuprofen Ketoprofen Naproxen Paracetamol Salicylic Acid Ciprofloxacin Sulfamethoxazole Trimethoprim Erythromycin Azithromycin Norfloxacin Levofloxacin Ofloxacin Atenolol Metoprolol Propranolol Diltiazem Caffeine Paraxanthine Carbamazepine Estriol 17-β Estradiol Estrone

Analgesic

3

Antiepileptic drugs Hormones

Stimulant

β-blockers

Antibiotics

PhACs

Classes

Table 1 Chemical profile of selected PhACs.

296.024 415.379 207.138 255.102 229.0851 152.0706 137.0249 332.141 254.0594 291.1452 734.5 749.519 320.1405 362.143 362.1511 267.1703 268.1907 260.1645 415.1686 194 181.07 237.102 287.1653 271.1704 269.1547

Mass to charge ratio (m/z) 4.2 10.6 4.9 4.5 4.2 9.4 3.49 6.38 5.7 6.8 8.8–8.9 8.74 6.34,8.76 6.24, 8.74 5.97, 9.28 9.6 9.6 9.45 8.06 10.4 8.5 13.9 10.54 10.46 10.34

pKa 0.70 1.19 3.97 3.12 3.18 0.46 1.19 0.4 0.89 0.91 3.06 4.02 0.46 −0.39 −0.39 0.16 2.15 −0.45 2.70 −0.07 1.17 2.45 2.45 4.01 3.13

Log Kow

(Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (“ PubChem Compound - NCBI,” 2018) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (“ PubChem Compound - NCBI,” 2018) (“ PubChem Compound - NCBI,” 2018; Papageorgiou et al., 2016; Santos et al., 2010; Tran et al., 2017) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Tran et al., 2017) (Batt and Aga, 2005; Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Tran et al., 2017) (“ PubChem Compound - NCBI,” 2018; Papageorgiou et al., 2016; Simazaki et al., 2015; Tran et al., 2017) (“ PubChem Compound - NCBI,” 2018) (“ PubChem Compound - NCBI,” 2018) (“PubChem Compound - NCBI,” 2018) (“ PubChem Compound - NCBI,” 2018) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Tran et al., 2017) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Tran et al., 2017) (“ PubChem Compound - NCBI,” 2018) (“ PubChem Compound - NCBI,” 2018) (Batt and Aga, 2005; Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Tran et al., 2017) (“ PubChem Compound - NCBI,” 2018) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Simazaki et al., 2015; Tran (“ PubChem Compound - NCBI,” 2018) (Behera et al., 2011; “PubChem Compound - NCBI,” 2018; Santos et al., 2010; Tran et al., 2017) (Behera et al., 2011; “ PubChem Compound - NCBI,” 2018; Santos et al., 2010; Tran et al., 2017)

Reference

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A. Majumder, et al.

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Fig. 1. Different sources and pathways of PhACs in aquatic environment (adapted from Nikolaou et al. (2007) and Sarmah et al. (2006).

America accounted for the maximum average concentration of analgesics (60 μg/L) in hospital effluent followed by Europe (17.7 μg/L). Paracetamol concentration was maximum in the United States (374 μg/ L) followed by ibuprofen (32.8 μg/L) (Oliveira et al., 2015a, 2015b). From Europe, the highest concentration of paracetamol in hospital effluent was found in Norway (325 μg/L) followed by Portugal (58 μg/L) (Langford and Thomas, 2009; Santos et al., 2013). Ketoprofen was also found in moderate concentrations (9.8 μg/L) in hospital effluent of Italy, which is higher than other European counties (Verlicchi et al., 2012). In surface water, the maximum average concentration of analgesics in surface water was in Africa (16.7 μg/L) followed by in North America (4 μg/L). High concentrations of ibuprofen (84.6 μg/L) and paracetamol (106.9 μg/L) were profound in the rivers of South Africa and Kenya, respectively (Matongo et al., 2015a, b; K'oreje et al., 2016). Salicylic acid, which is a metabolite of aspirin, was found in surface waters of the United States at a moderate concentration of 17 μg/L (Peng et al., 2014). Paracetamol, ibuprofen, and diclofenac were frequently detected in surface water of Asia, Europe, and North America with their concentration ranging from 13 ng/l to 6.4 μg/L (Tixier et al., 2003). Although these values are considerably less, prolonged intake or exposure of such contaminated surface water can kill microorganisms and also human beings. Antibiotics are prevalent in influent of WWTPs. The concentration of different antibiotics in the influent and effluent of the WWTPs depend primarily on the consumption pattern of these antibiotics by the people of that geographical region. It also depends on the daily socioeconomic conditions, water usage pattern, climatic conditions and the effectiveness of the treatment plants (Tran et al., 2017). Ciprofloxacin concentration of 31 mg/L was found in the industrial wastewater effluent in Tamil Nadu, India (Mutiyar and Mittal, 2014). This is the by far the highest reported concentration of any PhAC in any aqueous environment across the world. High concentrations of levofloxacin (196 μg/L), azithromycin (303 μg/L) and erythromycin (10 μg/L) were also found in India (Mutiyar and Mittal, 2014). The average concentration of antibiotics in influents of Africa (39 μg/L) and Asia (26.5 μg/L) was much higher than that of Europe (2.5 μg/L) and North America (1.51 μg/L). This indicates a higher consumption of antibiotics in the developing countries of Asia and Africa. This is a direct result of the lifestyle, climate, and low-income of people in these countries. People of India, Bangladesh, Pakistan, Thailand, China, and Kenya are

also been reported to be found in goldfish (Carassius auratus), mosquitofish (Gambusia holbrooki) and snails (Ebele et al., 2017). 2.2. Worldwide occurrence of PhACs in various water bodies Maximum pharmaceutical sales were observed in North America, followed by in Europe and Africa (Statista, 2018). Although the pharmaceutical consumption observed in the western countries was much higher than in Asian countries, the reported concentrations of various PhACs in different water bodies in Asia were at par or higher than those reported in North America and Europe. This shows that occurrence of pharmaceuticals in the aquatic environments depends on a number of factors like characteristics of water and soil of the region, climate, socioeconomic conditions of people, pharmacokinetics, treatment techniques available, etc. Significant concentrations of these PhACs have been found mostly in WWTP influent (Table S1), hospital effluent (Table S2), and surface water (Table S3). The only major difference is the concentration range in which they appear in the aforesaid sources. A detailed occurrence of major pharmaceuticals in WWTP influent, hospital effluent, and surface water are shown in Fig. 2. Analgesics are the most studied class of pharmaceuticals because of their high usage and availability in influents (Tran et al., 2017). Paracetamol, ibuprofen, and diclofenac are the most commonly occurring analgesics in influents. More than 60 μg/L and 50 μg/L of paracetamol were reported to be found in influent of WWTPs in Portugal and Canada, respectively (Guerra et al., 2014; Paíga et al., 2016). The maximum detected concentrations of ibuprofen, naproxen and diclofenac were 55.975 μg/L, 25 μg/L, and 22.628 μg/L, respectively found in Singapore, Canada, and South Africa, respectively (Guerra et al., 2014; Matongo et al., 2015a, 2015b; Tran and Gin, 2017). The average concentration of analgesics in WWTP influent were lower in Asia (4 μg/L) and North America (5.7 μg/L) as compared to Africa (39 μg/L) and Europe (43 μg/L). Analgesic consumption was found to be maximum in countries like Sweden and France. The per-capita consumption of analgesics was almost half in the USA as compared to the countries in Europe. This trend is reflected in the analgesic residuals in the wastewater of the respective countries (Diener et al., 2008). The concentrations of analgesics in hospital effluent followed a similar trend to that of the WWTP influent. Paracetamol, diclofenac, ibuprofen, ketoprofen, and naproxen were most frequently detected in hospital effluent. North 4

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Fig. 2. Worldwide occurrences of PhACs in WWTP influent (a), hospital effluents (b), and surface water (c). Data from: Table S1, S2 and S3.

more susceptible to viral fevers, diarrhea, and other febrile illness. Often, due to the lack of proper medical check-up antibiotics are consumed inappropriately. As a result, the non-metabolized parts of these antibiotics are more profound in the influents of Asia and Africa (Van Boeckel et al., 2014). Ciprofloxacin and sulfamethoxazole were the highest occurring antibiotics in the North America and Europe. More than 5.4 μg/L and 1.3 μg/L of ciprofloxacin were found in USA and Portugal, respectively while 3.1 μg/L and 2.6 μg/L of sulfamethoxazole

were found in USA and Canada, respectively (Guerra et al., 2014; Kostich et al., 2014; Mohapatra et al., 2016). The maximum average concentration of antibiotics was found to be the highest in Asia (27.9 μg/L). High concentrations of antibiotics in hospital effluent of Asia is directly related to the population and climate of the countries. High concentrations of ciprofloxacin (237 μg/L) and norfloxacin (29.6 μg/L) were found in hospital effluent of India (Diwan et al., 2010; Subedi et al., 2017). The occurrence of antibiotics in hospital effluent 5

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follows a similar trend to that of influents. The average concentrations of antibiotics were much lower in Europe (7.48 μg/L) and North America (4 μg/L) with sulfamethoxazole being the most frequently occurring antibiotic. Highest concentrations of sulfamethoxazole were detected in hospital effluent of Portugal, and the United States were 8.7 μg/L and 2.17 μg/L, respectively (Oliveira et al., 2015a, 2015b; Santos et al., 2013). Ciprofloxacin was found in moderate concentrations of 38.6 μg/L and 26 μg/L were also detected in hospital effluents of Portugal and Italy, respectively (Santos et al., 2013; Verlicchi et al., 2012). Antibiotics in surface water at most of the places were detected at concentrations less than 1 μg/L (Fig. 2). However, 6.5 mg/L of ciprofloxacin was found in Kazipeli Lake, Hyderabad, India (Mutiyar and Mittal, 2014). The maximum average concentration of antibiotics was the highest in Asia (48.4 μg/L) followed by in Africa (10.59 μg/L). Kenya's river water accounted for high concentrations of sulfamethoxazole (38.85 μg/L) and trimethoprim (6.95 μg/L) (K'oreje et al., 2016). Antibiotics in surface waters of Europe and North America were detected in concentrations lower than 500 ng/L. Researchers have reported that β-blockers are frequently detected in influent and effluent of WWTPs and hospital effluent because of their low biodegradability. Atenolol and metoprolol, two of the most frequently occurring β-blockers in influents and surface water have been reported in the range of 1 ng/L to 106 ng/L as depicted in Fig. 2(Kasprzyk-Hordern et al., 2008; Kostich et al., 2014; Li et al., 2016; Yang et al., 2017a, 2017b). Asia accounts for the maximum average concentration of β-blockers (45 μg/L) followed by the European countries (11.6 μg/L). Maximum concentrations of atenolol (0.3 mg/L) and metoprolol (0.95 mg/L) have been found in the influent of India (Mohapatra et al., 2016; Mutiyar and Mittal, 2014). Atenolol and metoprolol were the most commonly occurring beta-blockers in hospital effluent. The average concentration of βblockers in North America and Europe were 1.59 μg/L and 3.38 μg/L, respectively. Atenolol concentration of 6.6 μg/L in Italy was the highest reported value of β-blockers in Europe while, 3.54 μg/L of metoprolol detected in the United States was the maximum reported concentration in North America (Oliveira et al., 2015a, 2015b; Verlicchi et al., 2012). Propranolol was another β-Blocker detected in many hospital effluents of Europe and North America at concentrations varying from 10 ng/L to 25 μg/L (Langford and Thomas, 2009; Oliveira et al., 2015a, 2015b; Santos et al., 2013; Verlicchi et al., 2012). β-blockers were found in high concentrations in river waters of South Africa and Kenya. The concentration of Atenolol in surface water of South Africa was 30 μg/L (Agunbiade and Moodley, 2014). Indian surface waters accounted for 3.18 μg/L and 7 μg/L of atenolol and metoprolol. Low concentrations of propranolol and diltiazem were found in surface waters of Europe and North America. Among antiepileptic drugs, carbamazepine was the most occurring drug found in influent of all major countries of Asia, Europe, and North America. The average concentrations of these drugs were slightly higher in Europe (5.4 μg/L) than in Asia (3.48 μg/L) while North America had much lower concentrations (420 ng/L). Around 15 μg/L of carbamazepine was detected in Spain (Yang et al., 2017a, 2017b). Carbamazepine concentrations of 2.8 μg/L and 240 ng/L were detected in the hospital effluent of Norway and the United States (Langford and Thomas, 2009; Oliveira et al., 2015a, 2015b). It was also the most frequently occurring antiepileptic drug in surface waters across the world. It's concentrations ranged from 13 ng/L to 3.24 μg/L in South Africa (Balakrishna et al., 2017; K'oreje et al., 2016). Among stimulants, caffeine was preponderant and was found in high concentrations in North America (82 μg/L) and Europe (123 μg/L). This can be accounted for the high consumption rate of coffee and caffeine based products in these continents. The per capita coffee consumption (kg/per capita/year) of United States, Canada, Finland, Sweden, and Norway are 4.2, 6.1, 12, 8.2 and 9.9, respectively (Smith, 2018). These values are very high when compared to Asian and African countries. However, countries like India, Vietnam, and Indonesia contribute significantly to coffee exports. This can be the reason for the

high average concentration of caffeine in Asia as well (43 μg/L) (Smith, 2018). Caffeine was found in concentrations above 150 μg/L in the UK (Baker and Kasprzyk-Hordern, 2013). Caffeine was also found in high concentrations in Singapore (144 μg/L) and South Africa (33.2 μg/L) (Matongo et al., 2015a, b; Tran et al., 2014). The average concentrations of hormones in influent do not vary significantly from continents because it is mainly a constituent of domestic wastewater. North America has accounted for the maximum average concentration of hormones in influent (869 ng/L), which is slightly higher than the other continents. Estrone and 17-β Estradiol were found in South Africa at a concentration of 351 ng/L and 199 ng/L, respectively (Manickum and John, 2013). Hormones were also detected in low concentrations in hospital effluent of Iran, Korea, Belgium and Norway. Hormones were also detected in river waters of Asia, Africa, and North America. The concentration of estriol in Singapore (451 ng/L) was maximum among the other hormones occurring in surface waters. (Table S2). 3. Ecological impact The PhACs are persistent in the aquatic environment and can severely affect microorganisms and other living organisms present in the water on exposure (Bhatia et al., 2017). They enter the aquatic organisms and starts to go up the food chain resulting in biomagnification of these contaminants. Direct intake of these pollutants is also possible when human beings and animals drink the already contaminated water. Prolonged exposure to these PhACs among non-target species can lead to many kinds of anomalies (Zenker et al., 2014). 3.1. Evaluation of drinking water equivalent limit (DWEL) Since there are no proper guidelines regarding the drinking quality parameters for the pharmaceuticals, a drinking water equivalent limit (DWEL) for the respective drug was calculated based on certain governing parameters (Eq. (1)) (Benson et al., 2017; de Jesus Gaffney et al., 2014). The intention of toxicological assessment is to identify the maximum no-observed-adverse-effect-level which are based on animal, microorganisms, and human data. Reference doses for every pharmaceuticals are usually given in mg/day or μg/day. Since we are dealing with PhACs in the aquatic environment, the permissible limit should be defined in mg/L. An adjustment to the acceptable daily intake (ADI (mg/kg)) for the prescription drugs was accomplished by factoring the assumed weight of the consumers and the daily water intake of the consumers. Hence, DWEL was chosen to estimate the ecological risks as it takes into account the daily water intake of an average person, their body weights, which differ from region to region. The DWEL values of the PhACs have also been compared with the WWTP influent and hospital effluents because there may be a chance that the treatment plant at a particular region does not have the facilities to remove such micro-pollutants (de Jesus Gaffney et al., 2014).

DWEL =

ADI BW DWI

(1)

where, ADI is the acceptable daily intake (μg/kg-day), BW is the 50th percentile body weights (kg) of adults, DWI is the average daily water intake limit (L/day) for adults (de Jesus Gaffney et al., 2014). ADI indicates the limiting concentration of any particular drug which if not exceeded, will not have any adverse effect upon prolonged exposure. The values of ADI were primarily taken from various literature (EAEMP, 1995; de Jesus Gaffney et al., 2014; EPHC/NHMRC/NRMMC, 2008; Schwab et al., 2005; Snyder, 2008). For the PhACs whose ADI values were not available, no observed adverse effect limit (NOAEL) values were taken into the calculation of DWEL (Pfizer, 2006). The values of different BWs of adults from various countries were taken from BMC Public Health, and DWI (Daily Water Intake) was taken to be 3.7 l/day (Sawka, 2005; Walpole et al., 2012). The predicted no effect concentration for ecological toxicity to aquatic organisms (PNEC) and 6

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Table 2 Comparison of PNEC and calculated DWEL values with the reported concentration of selected PhACs.

calculated DWEL of the individual PhACs for various continents were used as the lower and upper benchmarks, respectively for comparing the maximum reported concentrations of the PhACs from that area. The calculated DWEL and PNEC are given in Table 2. The areas where the reported concentrations of the PhACs are less than the PNEC (if any) were marked with green. Areas with PhACs concentration greater than PNEC but lesser than DWEL were marked yellow while, areas with PhAC concentration higher than DWEL were marked with red. The green color indicates no threat from that particular PhAC upon water consumption, while yellow and red colors indicate moderate and high threats, respectively.

mechanisms of aquatic microorganisms. They have reportedly had adverse effects on the reproduction, locomotion, and metabolism of mussels. They can interfere with the endocrine system and cause disruption of homeostasis. Complex mixtures of PhACs and other organic contaminants can have synergetic toxic effects. A mixture of estradiol and 4-tert-nonylphenol led to a synergetic reaction and induced vitellogenin production in juvenile rainbow trout. A mixture of carbamazepine and clofibric acid had much more adverse effect on Daphnia Magna than the individual compounds at the same concentration (Ebele et al., 2017). Prolonged exposure to analgesics may overwhelm the ability of the enzymes to handle a substrate. As a result, there may be a disproportionate increase in plasma concentrations leading to intoxication (Boyer, 2012). This is how they are lethal to aquatic species. Prolonged exposure to antibiotics attack the tissues of the organisms and alter the metabolites produced during metabolism. These altered metabolites help in making bacteria more resistant and thus destroys the immune system. β-blockers prevent the interaction between catecholamine and β-adrenergic (G-protein-coupled receptors). This results in decreased calcium flow into the myocardial cells and decreases the blood pressure (Clarke et al., 2010). They are also known to hinder the cell regeneration (Haro et al., 2017; Yu et al., 2016). Hormones or sex

3.2. Impact on human, aquatic lives and microorganisms Prolonged exposure to these PhACs can lead to all kinds of anomalies among non-target species. Presence of PhACs in the aquatic environment in concentrations higher than that of the PNECs can harm aquatic life while prolonged intake of PhACs at a concentration higher than the DWEL can adversely affect the health of human beings. PhACs are designed to be highly active even at low doses and can affect metabolic and enzymatic 7

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Table 3 Ecotoxicological impacts of selected PhACs on non-target species. PhACs

Effects on non-target species

Reference

Diclofenac

Causes cytological changes in the liver, kidneys, and gills of fishes. Renal lesions. Reduces the hematocrit values of fishes. Exposed organisms are intoxicated due to increase in plasma concentrations. Postembryonic development among amphibians is hindered. Reproduction system of aquatic organisms is affected. Causes gastric ulceration, dyspepsia, bowel inflammation, mucosal damages and affects kidney, cardiovascular and central nervous system. Toxic to aquatic system. Lethal to aquatic species (Ceriodaphnia dubia, Thamnocephalus platyurus). Leads to the formation of hepatotoxic metabolites (N-acetyl-p-benzoquinone imine) and affects the liver. Immune system of humans and aquatic organisms are damaged. Accelerates the spread of antibiotic-resistant bacteria. Development of resistant pathogens is stimulated.

(Santos et al., 2010; Tran et al., 2017)

Codeine Ibuprofen

Ketoprofen Naproxen Paracetamol Ciprofloxacin Sulfamethoxazole Trimethoprim Erythromycin Azithromycin, Norfloxacin, Levofloxacin, Ofloxacin Atenolol Metoprolol Propranolol Carbamazepine Caffeine Paraxanthine Estrone, Estradiol, Estriol

Resists liver metabolic processes of fishes Inhibits the growth, metabolism of a species of cyanobacteria Resistant bacterial strains are formed which are harmful for terrestrial and aquatic organisms. Has toxic effect on microbes and destroys immune systems Growth of human embryonic stem cells are hindered. Highly toxic to aquatic organisms and plants. Causes cardiovascular and neural problems among humans Highly toxic to aquatic organisms. Cell regeneration process is hampered Growth of human embryonic cells is hindered. Causes anxiety and panic disorders among humans. May lead to endometrial, hepatocellular and colorectal cancer. Increases the diastolic blood pressure, free fatty acids and plasma epinephrine levels. It stimulates the sympathetic nerves and hampers the sympathetic nervous system. Severely affects the sexual and reproductive systems of, fish and humans. Lowers sperm count among males. Can cause birth defects, abnormal sexual development, and cancer. Affects the nervous system and the immune system.

hormones have the worst health impacts when consumed in large amounts (Hartmann et al., 2014). Many hormones and their metabolites can damage the DNA which can lead to cancer (Yasuda et al., 2017). Often incomplete degradation of certain PhACs like ibuprofen leads to the formation of metabolites (e.g. hydroxyl-ibuprofen and carboxy-ibuprofen) which may have similar or greater toxicity than the original compound (Ghauch et al., 2012). Also, if bromide ions are present in water, AOPs may lead to the formation of brominated by-products such as bromate and brominated organic compounds which are carcinogenic and genotoxic in nature (Aljundi, 2011). Ozonation of PhACs lead to the formation of by-products, which are at times as toxic or more toxic than the parent compound (Westerhoff et al., 2005). The various toxicological impacts on terrestrial and aquatic organisms upon prolonged intake of certain PhACs at a concentration higher than DWEL or PNEC (Table 2) are reported in Table 3.

Boyer (2012) (Ali et al., 2016; Santos et al., 2010; Tran et al., 2017; Xia and Lo, 2016) (Illés et al., 2014; Tran et al., 2017) (Santos et al., 2010; Tran et al., 2017) ( Li et al., 2017b; Santos et al., 2010; Tran et al., 2017) (Li et al., 2017b; Tran et al., 2017) (Nasuhoglu et al., 2011; Tran et al., 2017) (Jewell et al., 2016; Jiang et al., 2014; Tran et al., 2017; Zhang et al., 2011) Pérez et al. (2017) (Radosavljević et al., 2017; Yang et al., 2017a, 2017b) (Bhatia et al., 2017; Haro et al., 2017; Tran et al., 2017) (Soares et al., 2016; Tran et al., 2017) (Haro et al., 2017; Liu et al., 2013) (Tran et al., 2017; Xiong et al., 2016) (Ganzenko et al., 2015; Tran et al., 2017) Benowitz et al. (1995) (Hartmann et al., 2014; Pessoa et al., 2014; Tran et al., 2017)

treatment process of a WWTP in Korea with analgesic and hormones having a maximum removal efficiency of 28% (Behera et al., 2011). Secondary treatment methods involve aerobic or anaerobic processes to degrade the pollutants. As already mentioned, PhACs are highly toxic to the microorganisms, which prevent the microbial growth and hampers the process of degradation (Tran et al., 2017). Mohapatra et al. (2016) reported less than 10% removal efficiency for carbamazepine and metoprolol. Negative removal has also been observed for many pharmaceuticals like ibuprofen, ciprofloxacin, erythromycin, naproxen, atenolol, metoprolol, carbamazepine, and many other drugs (Mohapatra et al., 2016). The primary factor responsible for such a phenomenon can be because of the presence of these drugs in both their parent form and their conjugates. Upon biological treatment, these conjugates may undergo enzymatic reactions and form their original parent compound (Tran et al., 2017). Advanced or tertiary treatment involve processes like membrane filtration, adsorption by Granulated Activated Carbon and chlorination have been implemented in WWTPs (Yang et al., 2017a, 2017b). Although the main driving mechanism behind filtration is the sieving mechanism by which particles having sizes larger than the pore size are prevented from permeating, adsorption, and electrostatic interactions play a minor role. Due to the large pore sizes of microfiltration and ultrafiltration membranes, they are ineffective in removing PhACs. Nanofiltration has the potential to remove up to 97% of ionic contaminants, but some reports reveal removal of approximately only 60% of naproxen and diclofenac while carbamazepine had even less removal (Röhricht et al., 2009; Yang et al., 2017a, 2017b). Reverse Osmosis was reported to have effectively removed 99% of most PhACs. However, nanofiltration and reverse osmosis processes involve a huge amount of pressure and rejection rate which are the drawbacks to these processes. Granulated Activated Carbon (GAC) was reported to have removed

4. Existing WWTPs treatment processes and their removal efficiencies Conventional sewage treatment plants are generally not designed to tackle emerging persistent organic contaminants likes PhACs. They are mainly designed to remove organic carbonaceous compounds or phosphorus or nitrogenous substances (Tran et al., 2017). As a result, the basic processes like screening, sedimentation, flocculation, and biological treatment are not sufficient to remove PhACs(Mohapatra et al., 2016). Primary treatment methods like sedimentation and coagulation are not effective in removing PhACs because of their hydrophilic nature (Tran et al., 2017). Most of the pharmaceuticals have low log Kow values indicating that they are highly soluble in water and the chances of getting adsorbed on to the settling particles are very less. Most of the pharmaceuticals showed very little or no removal in the primary 8

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90–98% of most PhACs (Yang et al., 2017a, 2017b). However, only 43–64% of steroidal hormones were removed by GACs while carbamazepine was removed by only 17–23% (Grover et al., 2011). Furthermore, adsorption generates a large quantity of toxic sludge, which poses difficulty in sludge management.

catalysts with ozone (Miklos et al., 2018). 5.1.1. O3/H2O2 In this process, also known as peroxone process, H2O2 present in aqueous media forms peroxide anions that produces hydroxyl radicals upon reaction with ozone. Dose of ozone in peroxone process lies in the range of 1–20 mg/l and molar ratio of H2O2/O3 may be up to 0.5(Nöthe et al., 2009; Pisarenko et al., 2012). A small amount of H2O2 is also formed when O3 reacts with organic matter in water. Peroxone process shows enhanced removal of PhACs and even degradation of more recalcitrant organics. It also decreases the operation cost as lesser amount of ozone is consumed (Gerrity et al., 2011). Major limitation of this process is using the toxic hydrogen peroxide and has to be removed before disposing the effluent (Merényi et al., 2010a, 2010b; Nöthe et al., 2009). Alkaline pH (> 8), favors degradation of PhACs since there is an abundance of hydroxide ions thereby promoting the generation of OH% radicals by indirect ozonation. Alkaline pH makes ozonation a promising process which can be implemented to treat wastewater, however, precipitation of calcium carbonates may be of concern (Miklos et al., 2018).

5. Recent advances in advanced oxidation processes in the treatment of PhACs in an aqueous environment Since conventional treatment methods of WWTPs have not been able to effectively remove or degrade the PhACs present in the water, AOPs have gained immense popularity over the last decade in the field of water remediation (Do et al., 2019; Gora et al., 2018; Klavarioti et al., 2009; Li et al., 2018; Menapace et al., 2008; Zheng et al., 2016). AOPs are the processes where strong oxidation radicals (e.g. OH%, SO4%, O2%-, h+) capable of oxidizing and eliminating contaminants are generated “in-situ”. OH% radicals based AOPs like ozonation, photocatalysis, anodic oxidation, Fenton oxidation are the processes that have shown promising results in wastewater treatment where the contaminants are present in very low concentrations, as in the case of PhACs(Belver et al., 2017). Since the bond energy of the O–H bond (459 kJ/mol) is higher than the C–H (411 kJ/mol) bond, it can fetch one hydrogen atom from any organic compounds and degrade or break it (Buxton et al., 1988; Mckean, 1978; Navalon et al., 2010). Other radicals like (SO4%-) having higher oxidizing potential (2.5–3.1 V) than OH (1.9–2.8 V) can also be used as the main driving radical for degrading organic pollutants (Oh et al., 2016; Sahoo and Gupta, 2015; Wang and Wang, 2018a). Physical processes like ultrasound, microwave irradiation, electron beam, and plasma discharge have also gained immense popularity. The removal efficiency and time of these processes depend on various influencing parameters (e.g. pH, interfering ions, current intensity/density, electrolyte concentration, and catalyst dose). This review mainly encapsulates advanced oxidation studies conducted in lab-scale with synthetic wastewater. The experimental setup, removal efficiency, and experimental conditions for removal of PhACs by ozone-based AOPs have been presented in Table 4; photocatalysis, anodic oxidation, and sulfate activation in Table 8; Fenton based processes in Table 10 and physical processes in Table 11.

5.1.2. Ozonation with catalyst Catalytic ozonation is of two types: homogeneous and heterogeneous catalytic ozonation. In homogeneous catalytic ozonation, transition metal ions (Co, Cu, Fe, Mn, Pt, Ag, Au, etc.) are used as catalysts (Nawrocki and Kasprzyk-Hordern, 2010). Pines and Reckhow (2002) carried out homogeneous catalytic ozonation using Co(II) as the catalyst for the ozonation of oxalic acid. Firstly Co(II) forms complex with oxalate, then ozone oxidizes it to Co(III)-oxalate complex, and finally decomposed to Co(II) and oxalate radicals ( Pines and Reckhow, 2002). In homogeneous catalytic ozonation, metal oxide like TiO2, MnO2, ZnO, Al2O3, Fe-based, Ce-based etc are used as catalyst (Arfaeinia et al., 2016; Miklos et al., 2018; Nawrocki and Kasprzyk-Hordern, 2010). Both Homogeneous and heterogeneous catalytic ozonation shows lesser ozone consumption than ozonation alone but follows a more complex reaction pathways (Miklos et al., 2018; Wu et al., 2016). Catalytic ozonation shows complete removal of PhACs except for salicylic acid that may be due its recalcitrant nature (Vel Leitner et al., 1999). Catalytic ozonation has limited application due to recovery of catalyst and mechanism involved in the radical formation (Miklos et al., 2018). Mechanisms controlling the catalytic ozonation may be problematic, as the catalysts used in aqueous solutions will lead to competition between ozone, water and organic compounds for catalytic (adsorptive) active sites (Nawrocki and Kasprzyk-Hordern, 2010).

5.1. Ozone-based AOPs Ozone-based AOPs selectively oxidizes the organic matter having double bonds and higher electron density (Miklos et al., 2018; Rein Munter, 2001). Ozonation occurs in two ways: direct molecular ozone oxidation and/or indirect ozonation through ozone decomposition and the generation of hydroxyl radicals (OH%). Although this process shows good removal (> 90%) for PhACs, there are some limitations like incomplete mineralization, costly ozone consumption, etc. Saeid et al. (2018) have reported that ibuprofen shows 93% removal over 4 h ozonation, but on using Fe-based catalyst the rate of reaction and removal efficiency get enhanced and ultimately ozone consumption get reduced (Saeid et al., 2018). Reactions of ozone with organic compounds usually lead to the formation of aldehydes and carboxylic acids, which further do not react with ozone. As a result, complete mineralization of organic matter is not attained. Presence of bromide ions may lead to the formation of brominated byproducts such as bromate and brominated organic compounds which are carcinogenic and genotoxic in nature (Aljundi, 2011). Also, oxidative reactions with ozone are relatively slow and selective therefore modifications of ozonation process have been considered to overcome the limitations (Nawrocki and Kasprzyk-Hordern, 2010). In aqueous medium, ozone reacts with hydroxyl ions to form hydroxyl radicals which have 106 to 1012 times more oxidizing ability than ozone (Ribeiro et al., 2015). During ozonation, hydroxyl radicals are produced by means of activation methods or by reacting with the organic matter present in the water. Activation methods include integration of H2O2, UV irradiation, alkaline pH, and

5.2. Photocatalysis Photocatalysis is a process where solid semi-conductors having low band gap are excited by photons from a light source (e.g. UV, LED, Visible and Solar) (Gora et al., 2018; Zheng et al., 2016). If the photons have an energy greater than the band gap of the photocatalysts, electron-hole pairs are generated (Belver et al., 2017). The generated holes (h+) react with the water molecules to form hydroxyl radicals (OH%), which in turn reacts with the organic contaminants to degrade them to water, carbon dioxide and intermediate compounds (Leong et al., 2014; Wang et al., 2012) (Eq. (2) and Eq. (3)). Other non-toxic oxidants (e.g. O2%-, h+) present in the aqueous system and interfering parameters like (pH, catalyst dose, light source, and water matrix) may also have a significant role in the degradation of the PhACs(Debnath and Gupta, 2018; Wang et al., 2012) (Table 5). (2) (3)

9

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Table 4 Experimental conditions and removal efficiencies of PhACs in synthetic wastewater by Ozone-based process. AOPs/PhACs

Experimental conditions

Removal Efficiency (%)

Reference

100

Vogna et al. (2004)

Ibuprofen Ketoprofen

Initial Conc.: 1 mM; O3 dose: 1.0 × 10−3 M, pH:7, Time: 90 min for mineralization, 10 min, Sample volume 0.42 L Initial conc.: 10 mg/L; O3: 1000 mL/min; 240 min; pH:3; Sample volume 1100 mL Initial conc: 0.1 mM; O3: 4.94 mM for 60 min, pH 7.0, 25 °C; 300 mL volume

> 80 100

Naproxen

Initial concentration: 15 mg/l; O3: 0.22 mM, pH 3–7; 25 °C; 1L volume

100

Paracetamol Salicylic acid Ciprofloxacin Sulfamethoxazole Trimethoprim

pH 2.0 and 7.0; Initial Concentration = 5.0 mmol/L; T = 25 °C; Time: 120 min At pH 4 and in the presence of 1 mg/L of ozone; 10 min Initial Con. 15 mg/l; O3: 2.5 g/l for 75 min, pH 7, 27.5 °C Initial Concentration: 0.5 μmol; pH:8; O3: 0.2 mg/l Time: 15 min; in 15 min, Ozone dose of 5 mg/l

100 95 95 97–99 > 90

Erythromycin Norfloxacin Levofloxacin Ofloxacin Atenolol

Initial Concentration: 0.68 μM; O3: 6.8 μM for 2 min O3: 1.5 mg/l; Initial conc: 10 mg/l; 30 min time Initial concentration: 20 mg/l; O3: 20 mg/l; pH: 6.5; Time: 30 min Initial concentration: 22 mg/l; O3: 390 mL/min, pH 7.4, 25 °C; Time: 10.3min Initial conc: 100 mg/l; O3: 0.7 g/h for 20 min, pH 2,7,9, 25 °C; 1L volume

> 70 100 100 100 100

Metoprolol

Time: 15 min; Ozone dose of 5 mg/l; 500 mL volume

> 90

Propranolol Carbamazepine

Initial concentration:10 μM; pH 8.4; Time: 15 min, Ozone dose of 5 mg/l; 500 mL

> 90

Estriol 17-B Estradiol Estrone O3/catalyst Diclofenac

Initial concentration: 100 mg/l; pH:3–4; O3: 30 mL/min; Time: 90 min; 200 mL Initial concentration = 100 nmol L−1; T = 20 °C; contact time = 120 min; O3 = 1.5 mg L−1; 1.2 L Initial concentration: 1 mg/l; pH:6; O3: 20 mg/l, Time: 20 min; 250 mL volume

100 100 100

Saeid et al. (2018) (Illés et al., 2014; Wang and Wang, 2016) (Wang and Wang, 2016; Rosal et al., 2008) Andreozzi et al. (2003) Hu et al. (2016) DeWitte et al. (2008) ( Huber et al., 2003) (Luo et al., 2014; Sui et al., 2010) Luiz et al. (2010) Liu et al. (2012) Nasuhoglu et al. (2012) Carbajo et al. (2015) (Wang and Wang, 2016; Tay et al., 2011) (Luo et al., 2014; Sui et al., 2010) Benner and Ternes. (2011) (Luo et al., 2014; Sui et al., 2010) Ogata et al. (2011) Alum et al. (2004) Ben Fredj et al. (2017)

Ozone dose: 5.52 mg/l, flow rate of oxygen: 1.0 L/min, FSO/PMC dose: 800 mg/l, initial DCF concentration: 29.6 mg/l, Time: 20 min; temperature: 25 °C, initial pH: 7.0; 500 mL volume Initial conc.: 10 mg/L; T: 5° celcius degree celsius; 0.35 g/l H-Beta-25 zeolite catalyst; O3: 500 mL/min; 240 min; pH:3; 1100 mL Initial conc: 10 mg/l; TiO2+ O3; 10 min Time; 2 L volume MgO dosage: 2 g/L, O3 dosage: 1.8 mg O3/min, Initial concentration: 50 mg/L, pH: 5.4; 10 min time; 50 mL

100

Gao et al. (2017)

100

Saeid et al. (2018)

100 100

Initial conc.: 5 mM/l; pH: 7.2; Al2O3 Catalyst: 6g/46 mL of solution; O3: 0.3 mg; Time: 10 min; 2L reactor volume β-FeOOH/Al2O3 pH: 7; Catalyst dosage: 1.5 g/L; [O3]: 30 mg/L; flow rate: 12 L/h; Initial conc: 10 mg/L; 10 min; 1.2L reactor volume Sulfamethoxazole Initial conc: 50 mg/l; Ce-doped activated = 0.14 g L−1; pH 4.8; O3: 50g/m3; 30 min; 700 mL Azithromycin Initial conc.: 50 mg/l; pH 8.5; MgO catalyst dose 6 g/l; Time 6 min Norfloxacin Time 15 min; 0.1 g/l MnOx/SBA-15/O3; pH 5; 100 mg/h O3; 1.3 L reactor volume Atenolol Initial conc: 200 μg/l; Time: 120 min; Catalyst dose: 300 mg/l; O3: 380 ± 20 mg/l-hr; pH: 7; 1000 mL reactor volume Metoprolol Initial conc: 200 μg/l; Time: 120 min; Catalyst dose: 300 mg/l; O3: 380 ± 20 mg/l-hr; pH: 7; 1000 mL reactor volume Propranolol Initial conc: 200 μg/l; Time: 120 min; Catalyst dose: 300 mg/l; O3: 380 ± 20 mg/l-hr; pH: 7; 1000 mL reactor volume Carbamazepine Initial conc: 10 mg/l; TiO2+O3; 10 min Time; 2 L Rector volume 17-B Estradiol Initial concentration: 100 mg/l; 20 min time; O3: 20 mL/min; 1 g activated carbon; 200 mL volume O3/H2O2 Diclofenac O3: 5 mg/l; H2O2: 3.5 mg/l; Volume 40 L/min

65 100

Jankunaite et al. (2017) Mashayekh-Salehi et al. (2017) Vel Leitner et al. (1999) Yang et al. (2010)

100 100 100 100

Gonçalves et al. (2013) Arfaeinia et al. (2016) Chen et al. (2017) Wilde et al. (2014)

100

Wilde et al. (2014)

90

Wilde et al. (2014)

> 90 100

Jankunaite et al. (2017) Ogata et al. (2011)

99

Ibuprofen

83

(Gerrity et al., 2011; Luo et al., 2014) (Gerrity et al., 2011; Luo et al., 2014) Feng et al. (2015) De Witte et al. (2009)

O3 Diclofenac

Ibuprofen Ketoprofen Paracetamol Salicylic acid Ciprofloxacin

O3: 5 mg/l; H2O2: 3.5 mg/l; 40 L/min

Naproxen Ciprofloxacin

H2O2/O3 ratio = 0.5: 1; O3: 2 mg L-1 for 2 min; Initial concentration: 1 mg/l; 100 mL volume Initial concentration: 45.27 μM; O3: 2.5 g/l for 90 min, pH 7,27.5 °C, H2O2 concentration of 10 μmol/l; 1.75 L volume Sulfamethoxazole O3: 5 mg/l; H2O2: 3.5 mg/l; 40 L/min

5.2.1. Influence of catalyst dose and initial concentration Catalyst loading had a positive effect on photocatalysis of analgesics, antibiotics, β-blockers, antiepileptic, stimulants, and hormones but up to a certain extent. Increasing the dose beyond the optimal limit resulted in interference of transmittance of light, which resulted in reduced photocatalytic activity (Abellán et al., 2009; Gad-Allah et al., 2011; Mirzaei et al., 2018; Tao et al., 2015). TiO2 based photocatalysts were used to degrade β-blockers, atenolol, and metoprolol (Bhatia et al., 2017; Cavalcante et al., 2015; Ioannou et al., 2011; Ji et al., 2013; Romero et al., 2011; Ye et al., 2018). Bhatia et al. (2017) have reported

96–98 95 98

(Gerrity et al., 2011; Luo et al., 2014)

that catalyst loading had a positive effect on the degradation of atenolol but only up to 1.5 g/L, on further increasing the catalyst dose, the aqueous solution grew more turbid and hindered the transmittance of light and thereby hindering the photocatalytic activity (Bhatia et al., 2017). Cavalcante et al. (2015) have reported that 0.4 g/L catalyst dose can be chosen as the optimal dose for metoprolol removal because the degradation rate drastically decreased upon further increment of the photocatalyst dose. However, removal efficiency slightly improved when the loading was doubled because of adsorption by the powdered photocatalyst (Cavalcante et al., 2015). Most of the PhACs get adsorbed 10

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Table 5 Influencing parameters and responsible radicals in degradation of selected PhACs by photocatalysis.

onto the surface of the catalyst to some extent. PhACs with high log kow values tend to get adsorbed more because of their hydrophobic nature. Similarly high initial concentration led to a turbid solution, thereby reducing the intensity of light reaching the catalysts leading to the lesser generation of oxidizing agents. Hence at higher initial concentrations, degradation rate of the PhACs declined (Bhatia et al., 2017; Boukhatem et al., 2017; Chen et al., 2018b; Gad-Allah et al., 2011; Ioannou et al., 2011; Ye et al., 2018).

commonly used photocatalysts is the quantum yield for generation of hydroxyl radicals is very low. In TiO2 photocatalysis the quantum yield for generation of hydroxyl radicals was reported to be only 4% which indicates that the remaining 96% of the absorbed photons were wasted by conversion to heat (Sun Lizhong, 1996). The performance of the photocatalytic materials can be improved by overcoming barriers like recombination of electron and holes, reduction of band-gap, enhancing thermal and electrochemical stability and making them sensitive to visible lights. Metals like gold, silver, aluminum are used for surface doping of the catalyst, which prevents electron-hole recombination and energy loss (Debnath et al., 2019). Inorganic element doping of nitrogen and sulfur leads to reduced band-gap and make them sensitive to visible light. Carbonaceous materials provide thermal and electrochemical stability and helps in adsorption of the PhACs as well. Coating with polymeric materials and surfactants increase the light absorption capacity of the photocatalyst and sensitizes electrons on the catalytic surface (Debnath et al., 2019).

5.2.2. Influence of light source The source and wavelength of light is a major influencing parameter in photo-based degradation processes. UV based photocatalysis has shown better removal as compared to visible light based photocatalysis (solar radiation). Neppolian et al. (2002) observed degradation up to 33% for solar based photocatalysis while degradation under the same conditions with UV light was as high as 90% (Neppolian et al., 2002). Different UV lights of different wavelengths have a varying effect on degradation. The maximum quantum yield is for UV-C (wavelength: 200–290 nm) followed by UV-B (wavelength: 290–320 nm) and UV-A (wavelength 320–400 nm) (Rosario et al., 1979; Wang and Wang, 2018a). The quantum yield of visible light is lesser than the UV lights. As a result, the best degradation was achieved when UV-C was used as the light source. However, the type of light required for efficient photocatalysis depends completely on the catalyst used and its band-gap.

5.2.4. Role of oxidizing radicals Hydroxyl radicals played the most significant role in the degradation of β-blockers, antibiotics, hormones, and antiepileptic drugs (Cai and Hu, 2017; Doll and Frimmel, 2005, 2004; Gad-Allah et al., 2011; Ji et al., 2013; Kim et al., 2017; Sornalingam et al., 2018). However, Marques et al. (2013) and Ji et al. (2013) found O2% and h+ to be the major degrading radicals for analgesics and stimulants respectively, but they played second fiddle in degrading β-blockers (Ji et al., 2013; Marques et al., 2013). This can be explained by studying the frontier electron densities (FEDs) of highest occupied molecular orbital (HOMO), lowest unoccupied molecular orbital (LUMO) and point charges of the organic compounds. As per Frontier Orbital Theory, the electrophilic OH% radicals attack positions with higher FED2HOMO + FED2LUMO values, h+ attack positions with higher FED2HOMO value, while positions with most positive charge are attacked by O2%-( Wang et al., 2018b). Analgesics and stimulants may have more positions with higher FED2HOMO value and positive charges, making them reactive sites for h+ and O2%- radicals. On the other hand antibiotics, β-

5.2.3. Type of catalyst Photocatalysts are usually semiconductor-based materials that have a low band gap so that when light energy is absorbed by the materials, the electrons can easily jump from the valence band to the conduction band and thereby initiating the photocatalytic reactions. Semiconductor-based materials like titanium oxide (TiO2), Silica (SiO2), Zinc Oxide (ZnO), Tungsten oxide (WO3), zirconium oxide (ZrO2) have efficiently degraded PhACs. Other materials like carbon nanotubes, graphite-based materials, magnetite have also shown potential in degrading organic contaminants (Czech and Buda, 2015; Debnath et al., 2019; Laxma Reddy et al., 2017). However, a major drawback of many 11

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blockers, antiepileptic drugs, and many hormones have more positions with higher FED2HOMO + FED2LUMO values, making them susceptible to being attacked with OH% radicals.

Presence of phosphate ions also improved the photocatalytic degradation of metoprolol by getting adsorbed onto the catalyst surface and creating an electrostatic field. This electrostatic field promotes the separation of electron and holes (Ye et al., 2018). However, phosphate ions can also act as a scavenger for OH% radicals leading to poor degradation of analgesics like naproxen (Kanakaraju et al., 2015). Organic matters present in water matrix may get adsorbed on to active sites of the photocatalysts and can also act as scavengers for reacting radicals thus significantly reducing the rate of degradation of PhACs(Doll and Frimmel, 2005; Ioannidou et al., 2017; Sornalingam et al., 2018; Wang et al., 2018b; Ye et al., 2018). Chloride ions present in water reacts with OH% radicals to form secondary oxidants like Cl% and HOCl% with moderately high oxidizing potentials (2.40 V and 1.48 V, respectively) (Tan et al., 2017a, 2017b) (Eq.(4) and Eq. (5)). Also, Chloride radicals have a high affinity for holes (h+), which prevents recombination of electrons and holes thus facilitates production of OH% radicals (Kanakaraju et al., 2015; Salaeh et al., 2016). These secondary oxidants complimented the OH% radicals in increasing the degradation rate of antibiotics and analgesics (Kanakaraju et al., 2015; Mirzaei et al., 2018). However, Cl− ions acted as an OH% radicals scavenger and hampered degradation of hormones like estrone (Sornalingam et al., 2018).

5.2.5. Influence of pH of water matrix Apart from paracetamol and codeine, most of the analgesics have pKa (dissociation constant) values in the range of 4–5. The charge of a chemical compound in a solution of a particular pH may be estimated by knowing its pKa values. As per the Henderson-Hasselbalch equation, if the solution pH is less than the pKa value of the compound, it may get protonated and acquire a slight positive charge. Likewise, if the solution pH is higher than the pKa of a compound, it will remain un-protonated and may or may not acquire a net negative charge (Moore, 1985). High removal was observed at acidic and neutral pH. At acidic pH, analgesics exist as neutral species, and the photocatalyst gets positively charged, which facilitated some adsorption. Choina et al. (2013) used TiO2 as the photocatalyst to degrade ibuprofen, and the best removal was obtained at a pH of 7 (Choina et al., 2013). In case of ibuprofen, at neutral pH, they become negatively charged while the catalyst remained neutral. At such conditions, a certain amount of ibuprofen was adsorbed on to the surface of the catalyst (Choina et al., 2013). Since paracetamol has a high pKa value of 9.4, both the catalyst and paracetamol may be positively charged at lower pH. At basic pH, the catalyst acquires a negative charge while paracetamol remains positively charged. Also at basic pH production of OH% radicals get enhanced (Mirzaei et al., 2018). Thus paracetamol has the highest removal efficiency at a pH of 9 (Tao et al., 2015). Antibiotics had the best removal efficiency in acidic pH for various catalysts (Abellán et al., 2009; Cai and Hu, 2017; Chen et al., 2018b; Gad-Allah et al., 2011; Ioannidou et al., 2017; Mirzaei et al., 2018). The removal efficiency also depends on the point of zero charge (PZC) of the material. PZC indicates the pH at which the material has zero charge and at any pH lower than PZC would induce a net positive charge while any pH higher than the PZC would induce a net negative charge on the catalyst. Since pKa values of antibiotics and the PZC values of the most of the photocatalysts employed lies between 5 and 9, both the antibiotics and photocatalyst may develop positive charges in acidic medium. This shows that electrostatic attraction was not responsible for better removal. However, at such pH, some antibiotics (sulfa based) are known to have high light absorption and high photochemical reactivity leading to shorter half-lives and better degradation ( Boreen et al., 2004). It was found that near neutral pH (6–8) was found to be the ideal range for degrading β-blockers (Bhatia et al., 2017; Cavalcante et al., 2015; Ioannou et al., 2011; Ji et al., 2013; Ye et al., 2018). This is mainly because of high pKa values of atenolol and metoprolol indicating that they may get protonated and likely acquires a positive charge at such pH levels and gets attracted towards the negative surface of the photocatalyst (Ioannou et al., 2011; Ye et al., 2018). Similarly, neutral pH favors removal of carbamazepine, caffeine, estrone, and 17-β-estradiol because of their high pKa values. The stimulants, antiepileptic, hormones likely acquire a positive charge in an aqueous solution of neutral pH while catalysts with PZC values less than 6.5 or 7 remains neutrally charged or exhibits negative surface charge leading to adsorptive removal (Czech and Rubinowska, 2013; Elhalil et al., 2018; Kim et al., 2017; Martínez et al., 2011).

OH% + Cl− → HOCl%−

(4) (5)

5.3. Anodic oxidation (AO) Anodic oxidation oxidizes water with high O2 evolution overvoltage anodes (Pt, PbO2, SnO2, BDD) to form OH% radicals (Brillas et al., 2005; Chin et al., 2014; Oturan and Aaron, 2014; Wang et al., 2016). The hydroxyl radicals generated are physisorbed on to the surface of the anode leading to the formation of M (OH%), where M is the anode material (Eq. (6)). (6) (7) This heterogeneous hydroxyl radicals are more reactive than the OH % radicals and react with the contaminant to form carbon dioxide, water, inorganic ions and the anode material (Eq. (7)). Large quantities of other oxidants like peroxydisulfate (S2O82−), hypochlorous acid (HClO), biphosphate (P2O82−), peroxydicarbonate (C2O62−) are generated, due to the presence of sulfate ions, chlorine ions, phosphate ions, carbonate ions, etc. in the water matrix (Oturan and Aaron, 2014). Various parameters like pH, current density ,and influencing ions in the water matrix influence the rate of degradation of PhACs (Table 6). 5.3.1. Influence of current density The current density has a positive effect on the degradation and TOC removal. This is due to simultaneous generation of hydroxyl radicals and other active oxidants depending on the electrolyte. However, if the current density exceeds the limiting current density, oxygen evolution dominates the formation of active oxidants. Ciríaco et al. (2009) and Wang et al. (2016) observed that when Na2SO4 was used as the electrolyte for degradation of analgesics and antibiotics, the electrolyte got decomposed at high current density. Hence the degradation of analgesics and antibiotics were compromised at higher current densities (Ciríaco et al., 2009; Wang et al., 2016). Murugananthan et al. (2011) reported hindered atenolol degradation when higher current density lowered the formation of active chlorine from chlorine ions present in an electrolyte (sodium chloride). Since lesser active chlorine was available to oxidize the contaminants, the removal efficiency did not increase much with the increasing current densities (Murugananthan et al., 2011). However, there was no observed decline in degradation of antiepileptics, stimulants, and hormones with the increase of current density (Brocenschi et al., 2015; García-Gómez

5.2.6. Influence of interfering ions in a water matrix Ye et al. (2018) reported that the presence of HCO3− at a concentration of 50 mg/L had a net positive effect on degradation of metoprolol (Ye et al., 2018). It may be because HCO3− has a better scavenging effect for electron than hydroxyl radicals, which prevented recombination of OH% radicals. HCO3− reacts with hydroxyl radicals to form carbonate radicals having a high oxidizing potential of 1.78 V, which may help in degrading the β-blockers. However, higher concentration of CO32− scavenged more OH radicals leading to poor degradation of PhACs( Chen et al., 2018b; Ji et al., 2013; Sornalingam et al., 2018; Wang et al., 2017, 2012; Ye et al., 2018). 12

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Table 6 Influencing parameters and responsible radicals in degradation of selected PhACs by anodic oxidation.

et al., 2014; Indermuhle et al., 2013; Murugananthan et al., 2007; Shen et al., 2017). Higher current density is also not desirable as it significantly increases the electrical energy required thereby increasing the operation cost.

degradation (Pillai and Gupta, 2015; Wang et al., 2012). Similarly, when the electrolyte is Na2SO4, SO4%- is generated. Since SO4%- radicals have a higher redox potential than OH radicals, the degradation rate of PhACs is further accelerated.

5.3.2. Influence of pH of water matrix Analgesics tend to precipitate in acidic medium and are not easily dispersed to the surface of the electrodes. At higher pH, the analgesics start existing as acids, thus promoting their diffusion. As a result, at neutral pH, the ibuprofen, naproxen and diclofenac molecules can diffuse easily and reach the surface of the anode facilitating anodic oxidation (Brillas et al., 2010, 2005; Chin et al., 2014; Ciríaco et al., 2009; Murugananthan et al., 2011). Researchers have found that higher initial pH is preferred for degradation of antibiotics (Brinzila et al., 2012; Martín de Vidales et al., 2012a, 2012b; Sılva, 2017; Wang et al., 2016). This may be because of the abundance of OH radicals at higher pH (Wang et al., 2011a, 2011b). Also during degradation of antibiotics, aliphatic organic acids are formed, which lowers the pH of the solution inducing a positive charge on the antibiotics having pKa values (5.7–8.9), which may hinder their adsorption on the anode surface ( Wang et al., 2016). Murugananthan et al. (2011) reported that the degradation rate of β-Blocker was favored in acidic pH, but the maximum degradation was achieved at a pH of 6 (Murugananthan et al., 2011). Alkaline pH promoted the evolution of oxygen, which prevented the diffusion of the pollutants, thus preventing them from reaching the surface of the electrode. Inactive hydrogen peroxide (HO2−) was also formed at alkaline pH, which acted as a scavenger for OH% radicals. As a result, degradation was not favored at higher pH (Murugananthan et al., 2011). However, degradation of hormones showed an uncanny relation with pH. While 17-β estradiol, showed better removal at alkaline pH, estrone had near about the same degradation at acidic and neutral pH. Because of the high pKa value of 17-β estradiol, alkaline pH was required to ionize the molecules. Since anions are richer in electrons, they get readily attacked by electrophilic OH radicals (Murugananthan et al., 2007).

5.4. Degradation by activating persulfate (PS)/peroxymonosulfate (PMS) PS (S2O82−) and PMS (HSO5−) are oxidizers with relatively low oxidizing potential (2.01 V and 1.82 V, respectively). However, PS and PMS can be activated to generate SO4%- which have a much higher oxidizing potential (2.5–3.1 V) and can degrade organic contaminants at a rapid rate. PS and PMS can be activated to SO4%- by various activation processes. Thermal activation at a temperature above 50 °C causes a fission of O–O bonds in PS and PMS to form SO4%- ( Wang and Wang, 2018a). When UV light is applied in place of heat, similar fission of O–O bonds occur leading to the formation of SO4%-. It was found that the quantum yield of SO4%- (1.4) under UV light at 248 nm wavelength was maximum for PS while it was 0.12 for PMS. UV light of higher wavelength also had a decreased quantum yield of SO4%- (Herrmann, 2007) (Eq. (8) and Eq. (9)). (8) (9) Activation can also be achieved by using metals (Cu, Fe, Co, Ti, Bi, etc.) and their oxides. The activation of PS and PMS primarily depends on the redox potential of the metal (Wang and Wang, 2018a) (Eq. (10) and Eq. (11)). (10)

(11) Other modes of activation are alkaline activation, carbonaceousbased materials activation, ultrasonic activation, radiation activation, and hybrid activation systems ( Wang and Wang, 2018a). Degradation using SO4% also depend on the pH of the system, method of oxidation, interfering ions, and dose of PS/PMS (Table 7).

5.3.3. Influence of electrolyte type and concentration The electrolyte also played a significant role in the degradation of estrone. Researchers have reported that acidic pH favors the generation of active chlorine radicals (HOCl%, ClO−) in the presence of an electrolyte (NaCl), which increases the rate of degradation (Brocenschi et al., 2015; Pillai and Gupta, 2016). However, a higher concentration of Cl− scavenges OH% radicals, which leads to a decreased rate of

5.4.1. Influence of activation method of sulfate radicals Although both PMS and PS are used to generate SO4%-, PS holds a significant advantage over its counterpart. Due to the difference in the chemical structure of the two compounds, activation of PMS is tougher

13

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Table 7 Influencing parameters and responsible radicals in degradation of selected PhACs by activation of PS/PMS.

than PS. PS has bonds of O–O (bond energy: 142 kJ/mol) while PMS has O–H bonds (bond energy: 459 kJ/mol), making it harder to dissociate under thermal activation (Huheey and Cottrell, 1958). The number of SO4%- generated increases with the increase in dose leading to better degradation ( Chen et al., 2018a; Ghauch et al., 2012; Liu et al., 2013b, 2013a; Lu et al., 2017; Mahdi-Ahmed and Chiron, 2014; Tang et al., 2017). PS was activated via thermal, UV-light and metal oxides to degrade sulfamethazine, an antibiotic. Wang and Wang (2018a, 2018b), have elucidated that the time required to degrade for thermally activated PS was the most (6 h) while UV-light took the least time (2h) and metal oxides took 3h to degrade. Similarly, ciprofloxacin was degraded in twice the time when PS was activated thermally as compared to activation of PS via UV-light (Wang and Wang, 2018a). Researchers have reported that Metoprolol (β-Blocker) can be degraded in 1h when Cu/Fe2O4 was used for activation, while atenolol (β-Blocker) took 0.5h in presence of UV-light (Chen et al., 2018a; Wang and Wang, 2018a). Hence it can be inferred that thermal activation is the slowest of the processes followed by metal-oxide activation, while UV-light based activation is the fastest. For photo-activation of PS and PMS, UV-C, having the highest quantum yield fetched the best results (Wang and Wang, 2018a). However, highly turbid solutions can hinder transmittance of light and inhibits the activation.

a pH of 7 and 9 ( Chen et al., 2018a; Liu et al., 2013b, 2013a). Wang and Wang (2018a, 2018b) has elucidated that PMS is stable at pH less than 6 and at a pH of 12 ( Wang and Wang, 2018a). It was also reported that at pH 9, HSO5− dissociates to form SO52− which leads to the formation of singlet oxygen (1O2). 1O2 is a strong oxidant which attacks selective electron-rich compounds like amine which is a functional group present in β-blockers (Mehvar and Brocks, 2001; Wang and Wang, 2018a). However, at pH > 9.3, SO4% got converted to OH% radicals. Since the oxidizing potential of the SO4% were greater, there was a drop in degradation rate ( Liu et al., 2013b). Deng et al. (2013) reported that degradation of carbamazepine decreased with the increase in pH because of the recombination of generated OH% radicals at higher pH and SO4%- to form HSO4− and oxygen (Deng et al., 2013). Maximum degradation for caffeine was observed at neutral pH. At acidic pH, SO4%were not generated while at alkaline pH formation of metal hydroxide prevented the activation of PMS( Guo et al., 2015b). 5.4.3. Influence of interfering ions in a water matrix Researchers have found that SO4%- is the dominant radical behind the degradation of PhACs because of its higher redox potential while OH% radicals played a less significant role in degradation ( Chen et al., 2018a; Dulova et al., 2017; Ghauch et al., 2012; Liu et al., 2013b; Oh et al., 2017; Tang et al., 2017). The role of the water matrix and its effect on degradation depends directly on their scavenging or oxidizing property. It has been reported that the presence of HCO3− ions in low concentrations enhanced degradation of diclofenac and paracetamol ( Liu et al., 2013b; Lu et al., 2017; Tang et al., 2017). This was due to the formation of reactive carbonate species having high oxidizing potential (1.63 V), which took part in degradation (Lu et al., 2017). Lu et al. (2017) further reported that on increasing the dose beyond a certain point (25 mM) the SO4%- and OH% radicals got scavenged by the HCO3− and CO32− ions (Lu et al., 2017). As a result, for most of the PhACs, HCO3− and CO32− had a negative effect on degradation ( Guo et al., 2015b; Ji et al., 2016, 2014; Liu et al., 2013a; Mahdi-Ahmed and Chiron, 2014; Mehvar and Brocks, 2001; Oh et al., 2017). Similarly, researchers have found that organic matters had a negative effect on degradation because of its radical scavenging property and ability to decrease UV transmittance (Deng et al., 2013; Ji et al., 2016, 2014; Liu

5.4.2. Influence of pH of water matrix Optimum pH for degradation of analgesics and antibiotics was 7–9 (Ghauch et al., 2012; Mahdi-Ahmed and Chiron, 2014; Oh et al., 2017; Tang et al., 2017). However, Dulova et al. (2017) reported that naproxen had the worst degradation at pH 9, but that was because of the iron-based catalyst used for activation. At neutral or alkaline pH, generation of excessive hydroxyl ions led to the formation of excessive mixed chelation resulting in a decrease in availability of Fe2+ required for activation (Dulova et al., 2017). Lu et al. (2017) have reported that diclofenac showed the best degradation at a pH of 11 suggesting the generated OH radicals also played a significant role in its degradation (Lu et al., 2017). Trimethoprim showed the best degradation at a pH of 3 because acidic pH promoted electrostatic interactions between the protonated pyrimidine ring and SO4%-(Ji et al., 2016). Similarly, researchers found that β-blockers showed maximum degradation between

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Table 8 Experimental conditions and removal efficiencies of PhACs in synthetic wastewater by Photocatalysis, activated PS/PMS and anodic oxidation. AOPs/PhACs Photocatalysis Ciprofloxacin Erythromycin Trimethoprim Tetracycline Sulfamethoxazole Paracetamol Ibuprofen Naproxen Diclofenac Ketoprofen Atenolol Metoprolol Caffeine Carbamazepine

Experimental Condition

Removal Efficiency (%)

References

Dose:1.5 g/l Degussa P25; UV 254 nm; 125W; Volume:150 mL; pH: 7; Temp: 26 °C Time: 60 min; [Co]:3.3 × 104 μg/l Dose: 0.01 g Degussa P25, UV light; neutral pH; Time: 90 min, [Co]: 3 × 105 μg/l Dose: 2 × 10−4M Na4W10O32; pH: 6; Time: 200 min; [Co]:1 × 104 μg/l; UV-300 nm; 15W; Volume: 30 mL Time: 80 min; 300W Xe lamp 420 nm; Volume 100 mL; Dose: 500 mg/l Degussa P25; UVA; 9W; Time: 120 min; pH:4; [Co]: 1 × 104 μg/l; Volume 350 mL Dose: 0.5 g/l TiO2; Time: 360 min; [Co]:3.5 × 104 μg/l Dose: 0.05 g/l ZnO; Hg lamp 150W; pH: 7; Temp: 20 °C; Time: 60 min; [Co]=1 × 103 μg/l Dose: 0.1 g/l TiO2; UV lamp; 47W; pH: 3; Temp: 24–29 °C; Time: 180 min; [Co]: 3 × 104 μg/l; Volume 400 mL Dose: 0.05 g/l ZnO; Hg lamp 150W; pH: 7; Temp: 20 °C; Time: 60 min; [Co]: 1 × 103 μg/l Dose: 1 g/l P25 TiO2; Hg lamp 3W; pH: 4.6; Temp: 20 °C; Time: 240–360 min; [Co]: 1.5 × 104 μg/l; Volume 800 mL Dose: 2 × 10−4M Na4W10O32; pH: 6; Time: 45 min; [Co]: 1 × 104 μg/l, UV 300 nm; 15W; Volume 30 mL Dose: 0.4 g/l TiO2; UV lamp 273 nm; Time: 218 min; [Co]: 5 × 104 μg/l TiO2 nanotube; pH: 6; Time: 240 min; [Co]: 1.5 × 104 μg/l Dose: 0.2 g/l Cobalt ferrite nano-composite; Hg UV lamp 254 nm; 9W; pH: 2.5; Time: 100 min; [Co]: 5 × 103 μg/l; Volume 500 mL TiO2; UV lamp 254 nm; pH: 7; 700W; Time: 280 min; [Co]: 5 × 104 μg/l; Volume 300 mL Dose: 0.4 g/l TiO2 Degussa P25; UVC; Temp: 22–24 °C; pH: 4.6; Time: 200 min, [Co]: 1 × 105 μg/l

99

An et al. (2010)

100 90

( Ambrosetti et al., 2015) Molinari et al. (2017)

88 100 95 60 100

Zhu et al. (2016) Xekoukoulotakis et al. (2011) Rimoldi et al. (2017) Bohdziewicz et al. (2016) Kanakaraju et al. (2015)

68 100

Bohdziewicz et al. (2016) (Martínez et al., 2013)

> 95

Molinari et al. (2017)

90 50–60 > 96

Chen et al. (2018a, 2018b) Arfanis et al. (2017) He et al. (2017b)

80 100

Czech and Rubinowska (2013) Li Puma et al. (2010)

100 94.5 > 90 100

Mahdi-Ahmed and Chiron (2014) ( Li et al., 2017a) Zhang et al. (2016) Liu et al. (2017)

100 97

(Tang et al., 2017) Gong et al. (2017)

100 85

Dulova et al. (2017) Lu et al. (2017)

100 100

Feng et al. (2017) (Liu et al., 2013)

68.7 100 89.4

Romero et al. (2015) Rivas et al. (2012) Wang and Zhou (2016)

> 90

Liu et al. (2016c)

99.5

Deng et al. (2016)

99.96 100

Menapace et al. (2008) González et al. (2011)

95.2 98

Liang et al. (2018) Brillas et al. (2005)

99.32

Pérez et al. (2010)

97

Brillas et al. (2010)

100

Domínguez et al. (2010)

100

Sirés et al. (2010)

100 100

(Martín de Vidales et al., 2012a, 2012b) Indermuhle et al. (2013)

> 88

García-Gómez et al. (2014)

98

Brocenschi et al. (2015)

100

Murugananthan et al. (2007)

Estrone 17-β estradiol PS/PMS Activation Ciprofloxacin Dose: 1 mM Oxone; UVC; 15W; pH: 7; Time: 60 min; [Co]: 16567.3 μg/l; Volume 250 mL Erythromycin Dose: molar ratio nZVI/Persulfate: 4; pH: 5–7; Time: 250 min; Temp: 25 °C; [Co]: 1 × 103 μg/l Trimethoprim UV/peroxydisulfate; pH: 9; [Co]: 2.9 × 104 μg/l Sulfamethoxazole Dose: 1 mM Potassium perdisulfate; UV lamp; 10 kW; pH: 8; Time: 60 min; [Co]: 5 × 103 μg/l; Volume 100 mL Paracetamol Dose: 0.2 g/l MnFe2O4 and oxone; pH: 7; Time: 120 min; [Co]: 1 × 104 μg/l Ibuprofen Dose:0.25 mM [Fe2+/Oxone, Fe2+]; UV 300 nm; 35W,21W,24W; Volume 200 mL; pH: 3.68; Temp: 25 °C; Time: 20 min; [Co]: 1 × 104 μg/l Naproxen Dose: 1:10:10 M ratio NPX/H2O2/S2O82−; pH:3; Time: 180 min; [Co]: 17269.73 μg/l Diclofenac Dose:1 mM Persulfate; UV2-54 nm; 75W; Alkaline pH; Time: 60 min; Temp: 24 °C; [Co]: 9 × 104 μg/l; Volume 100 mL Ketoprofen Dose: 2 mM Persulfate; UV lamp; pH: 5; Temp: 40–70 °C; Time: 60 min; [Co]: 2.5 × 103 μg/l Atenolol Potassium peroxodisulfate; UV lamp 10W; pH: 7; Temp: 23 °C; Time: 30 min; [Co]: 5.3 × 103 μg/l; Volume 1.5L Metoprolol Dose: 0.5 g/L CuFe2O4 and 2 mM Na2SO3; pH: 11; Temp: 25 °C; Time: 60 min; [Co]: 2673 μg/l Caffeine Dose: 0.05 mM Co(II) and Oxone; pH: 2.1; Temp: 20 °C; Time: 120 min; [Co]: 2.35 × 107 μg/l Carbamazepine Dose: 1 mmol/l persulfate; Ultrasound radiation 50 W; pH: 3; Time: 120 min; Temp: 50 °C; [Co]: 5.1 × 103 μg/l; Volume 100 mL. Estrone 17-β estradiol Dose: E2:Persulfate= 1:10; Visible light radiation; Temp: 25 °C; pH: 3; Time: 60 min, [Co]: 5 × 103 μg/l Anodic Oxidation Ciprofloxacin SnO2–Sb/Ti electrode; 150W; Electrolyte: 0.05 mol/L Na2SO4; J: 30 mA/cm2; pH: 5.4; Time: 120min; Temp: 30 °C; [Co]: 5 × 104 μg/l; Volume 250 mL Erythromycin BDD anode, Electrolyte: 1 mol/l NaHSO4; Current density: 37.5 mA/cm2; [Co]: 50 μg/l Trimethoprim BDD anode; Electrolyte: 0.49 mol/l Na2SO4; Current density: 207 mA/cm2, pH: 3; Temp: 25 °C; [Co]: 5 × 104 μg/l Tetracycline pH: 6; Time: 300 min; Volume 100 mL Paracetamol BDD anode; pH: 2–12; Time: 90 min; Temp: 35 °C; Current intensity: 450 mA; [Co]: 1.57 × 105 μg/l Naproxen BDD/Si anode and stainless steel cathode; Current density: 100 A/m2; Time: 60 min; [Co]: 9.223 μg/l; 42W; Volume 2L Diclofenac BDD/stainless steel cells; Electrolyte: 0.05 M Na2SO4; Current intensity: 300 mA; Time: 360 min; pH: 6.5; [Co]: 1.75 × 105 μg/l Ketoprofen BDD electrodes; Electrolyte: 0.5 M Na2SO4; Current density: 235 mA/cm2; pH: 3.99; Temp: 25 °C; Time: 120 min; [Co]: 5 × 104 μg/l Atenolol BDD anode and carbon felt cathode; Electrolyte: 0.05 mol/l Na2SO4; Current intensity: 300 mA; pH: 3; .Temp: 25 °C; Time: 600 min; [Co]: 4 × 104 μg/l Metoprolol BDD anode, stainless steel cathode; Electrolyte: Na2SO4; Current density:15mA/cm2; Time: 240 min; [Co]: 100–105 μg/l Caffeine BDD anode and stainless steel cathode; Electrolyte: 0.035 M Na2SO4; Current density: 30 mA/ cm2; Temp: 25 °C; Time: 60 min; [Co]: 1 × 105 μg/l; 200W; Volume 21.4L Carbamazepine Ti/PbO2 circular anode; Electrolyte: 0.4 g/l Na2SO4; Time: 101 min; Temp: 25 °C; Current: 1.37 A; [Co]: 1 × 104 μg/l Estrone BDD anode; Electrolyte: 0.1 mol/l Na2SO4; Current density: 10 mA/cm2; pH ≤ 7; Time: 30 min; [Co]: 500 μg/l 17-β estradiol Si/BDD anode; Electrlyte: 0.1 M Na2SO4; Current density: 25 mA/cm2; pH: 6; Temp: 25 °C; Time: 60 min; [Co]: 500 μg/l

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et al., 2013b, 2013a; Lu et al., 2017; Mahdi-Ahmed and Chiron, 2014; Tang et al., 2017). However, Tan et al. (2017a, 2017b) reported that catalytic activation of PMS for removal of paracetamol followed a slightly different trend as degradation improved when the concentration of humic acid was increased up to 10 mM. Catalytic reaction with humic acid may be the driving factor behind this anomaly (Tang et al., 2017). When chloride ions are present in the system, SO4%- and OH% radicals react with it to form Cl% according to Eq. (4), Eq. (5), Eq. (12), and ) (Lu et al., 2017; Tang et al., 2017). Cl% with its relatively high oxidizing potential assisted in degrading diclofenac when the concentration of chloride ions was low. Also, the chloride ions provide an acidic medium, which helps in generation of more SO4%(Lu et al., 2017). However, at a higher dose more and more SO4% would be converted to less reactive Cl%, thus bringing down the rate of degradation (Deng et al., 2013; Ji et al., 2014; Liu et al., 2013a; Mahdi-Ahmed and Chiron, 2014; Tang et al., 2017) (see Table 8). (12)

acidic pH ideal for Fenton based processes. Also, low ferrous hydroxide production in acidic pH rendered less consumption of ferrous ion (Badawy et al., 2006; Velásquez et al., 2014). It was found that the complete removal of antibiotic at pH 3 while 80% removal at the neutral condition. A similar trend was reported in case of diclofenac and paracetamol (Alalm et al., 2015; Oturan and Aaron, 2014). Houtman (2010) observed that the initial increase in temperature led to an increased degradation rate, but on further increment beyond 40 °C inhibited degradation rate (Hartmann et al., 2010). This may be because the H2O2 decomposes to water and oxygen at a higher temperature. AOPs are usually carried out at room temperature (20–27 °C) ( Guo et al., 2015a; Hartmann et al., 2010). 5.5.1.2. Influence of initial concentration and dose. High initial concentration would lead to higher production of intermediates which in turn would consume the OH% radicals leading to decreased degradation and increased toxicity (He and Zhou, 2017). High dose of H2O2 leads to auto-decomposition to H2O and O2. It leads to reduced degradation/mineralization due to loss of OH radicals (Badawy et al., 2006; Domínguez et al., 2012; Elmolla and Chaudhuri, 2009). High dose of H2O2 is also undesirable because it is costly, non-recoverable, and has to be removed before discharging the effluent in water bodies because of its toxic nature (Bautista et al., 2010; Zazo et al., 2011). Catalyst dose (i.e. ferrous ion or ferric ion) had a positive effect on degradation, but a very high dose of the catalyst is not desirable due to scavenging of OH% radicals and formation of turbid water (Alalm et al., 2015; Tekin et al., 2006). Ifelebuegu and Ezenwa (2011) have reported that an increase in dose of ferrous ion above 50 mg/L, decreased the removal efficiency (Ifelebuegu and Ezenwa, 2011).

(13) 5.5. Fenton Process Fenton process is used for the degradation of organic contaminants in the presence of hydroxyl radicals originating from the reaction of hydrogen peroxide and Fe+2/Fe+3 under acidic condition (Moreira et al., 2017; Shemer et al., 2006). In presence of Fenton reagents, organic compounds are converted into carbon dioxide, water, and intermediate byproducts that may be toxic in nature. Fenton processes can be classified into two categories based on an in-situ generation or external addition of H2O2. Consumed catalyst and OH% radicals in these processes are usually regenerated by applying photo-radiation and electrolytic forces (Qiang et al., 2003). This is a significant advantage over simple catalytic reactions, where the reactivity of catalyst gets exhausted after a consecutive run due to leaching and exhaustion of catalytic material. Hydroxyl radicals are the primary responsible radicals for degradation in Fenton processes. The influence of process parameters and water matrix in degradation of PhACs of various Fenton processes are showed in Table 9.

5.5.1.3. Influence of ions. Presence of organic matters increasingly reduces the removal of PhACs due to its filter action and increased the consumption of the excess OH% radicals (Legrini et al., 1993; Li et al., 2015). However, Cardoza et al. (2005) reported that the removal rate of ciprofloxacin increased in the presence of organic matters. This may be due to the high reactivity of these PhACs in presence of organic matters and adsorption on to organic matters (Cardoza et al., 2005). High carbonate and bicarbonate concentrations may reduce the removal efficiency of PhACs due to scavenging of OH% radicals (Bautitz and Nogueira, 2007). The presence of inorganic ions such as chloride (Cl−), sulfate (SO42−), dihydrogen phosphate (H2PO4−) and hydrogen phosphate (HPO42−) react with OH% radicals to form anionic radicals, which are not efficient to remove PhACs(Sirtori et al., 2009). This significantly reduces the degradation of PhACs. Cl− will have less effect on removal than phosphate (PO43−), because it forms HOCl radicals, which has some redox potential, but less than that of OH% radicals (Mirzaei et al., 2017; Sirtori et al., 2009). Mirzaei et al. (2017) reported that the presence of PO43− completely inhibits the Fenton process and reduces the mineralization due to the formation of ironphosphate complexes (Mirzaei et al., 2017). SO42− has a less detrimental effect than Cl− and PO43− because of its high redox potential, which oxidizes the PhACs to some extent. In the case of antibiotics and antiepileptic drugs, SO42− increases the removal efficiency because of its high oxidizing potential (Ahmed and Chiron, 2014; Ji et al., 2014; Wang and Wang, 2018b). Often intermediate ions like nitrate (NO3−), ammonium (NH4+), and carboxyl ions are generated, which may lead to undesirable consumption of OH% radicals and resulting in reduced degradation of PhACs.

5.5.1. Conventional Fenton process In conventional Fenton process, PhACs are degraded and mineralized by the OH% radicals generated from the reaction of Fe+2/Fe+3 and hydrogen peroxide as shown in Eq. (14) and ) (Moreira et al., 2017; Shemer et al., 2006). (14) (15) Although conventional Fenton process is efficient in removing PhACs, it involves certain disadvantages like unpremeditated consumption of OH% radicals by H2O2 and Fe2+. Other significant disadvantages are consumption of H2O2 and production of ferric ions, which prevents recycling of Fe2+ ions. Also, after treatment using the Fenton process, acidification or neutralization of alkaline water is required due to the presence of hydroxyl compounds in it (Mirzaei et al., 2017). The efficiency of Fenton process depends on various influencing parameters such as reaction time, pH, temperature, the initial concentration of PhACs, H2O2 dose, catalyst dose, organic and inorganic ions, and reusability of catalyst (Ahmed et al., 2017; Mirzaei et al., 2017; Moreira et al., 2017).

5.5.2. Photo Fenton process In photo Fenton process, consumed Fe2+ ions present in the form of Fe3+ ions [Fe(OH)2+ and Fe(OH)2+] absorb light in the UV/visible region and undergoes photo-reduction to generate OH radicals and

5.5.1.1. Influence of pH and temperature. Researchers have found that acidic pH (3–4) favors a higher decomposition rate of H2O2,making

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Fe2+ as shown in Eq. (16)(Mirzaei et al., 2017; Moreira et al., 2017).

are available on electrolyte dose or concentration variation. High dose of the electrolyte is not desirable as it has an insignificant role in removal (Moreira et al., 2017). High concentration of electrolytes also generates more inorganic ions that may scavenge OH% radicals.

(16) In photo Fenton process, degradation efficiency is increased due to the increase in generation rate of OH% radicals under photo-radiation. Simultaneously, Fe3+ ions act as light absorbing species and get converted to Fe2+ and OH% radicals (Mirzaei et al., 2017). In photo Fenton process, less amount of sludge is produced as Fe2+ is further recovered from the leachate in presence of photo-radiation (Hermosilla et al., 2009). Effect of influencing parameters on degradation is same in photo Fenton process as in case of conventional Fenton process. This process is characterized by a significant increase in degradation rate and an increase in the life of catalyst due to its recovery.

5.5.3.2. Influence of electrode materials. In EF process, electrodes (cathode and anode) are usually of different materials. Cathodes are mostly carbon-based materials such as carbon-felt, air diffused electrodes whereas anodes are mostly metal and few carbon-based. Ruthenium oxide (RuO2), iridium oxide (IrO2), platinum (Pt), graphite are examples of active anodes and have potentials for O2 evolution at 1.8 V@ standard hydrogen electrode (SHE). Lead dioxide (PbO2), tin dioxide (SnO2), boron-doped diamond (BDD) and titanium dioxide (TiO2) electrodes are considered as non-reactive electrodes, having potentials for O2 evolution anode 1.7–2.6 V@ SHE(Moreira et al., 2017). BDD is widely used anode because of its non-reactive nature. It has also combined characteristics of substrates including Si, Ti, Nb, B, sp3/sp2 ratio and its unique layer thickness makes it suitable to oxidize any organic materials (Brillas and Martínez-Huitle, 2011; Moreira et al., 2017). High O2-overpotential of BDD is also beneficial for more H2O2 production (Ganzenko et al., 2015).

5.5.2.1. Influence of photo-irradiation and initial PhACs concentration. Degradation rate increases with the increase in radiation intensity of the UV/visible light used (Mirzaei et al., 2017). The use of photo-radiation (UV light) of lower wavelengths led to higher degradation rate, higher mineralization and near neutral operating pH which are significant advantages over conventional Fenton process (Aleksić et al., 2010; Wang et al., 2016a). Photoradiation also causes regeneration of Fe+2 and OH% radicals from the intermediate compounds. Higher initial concentration will cause a turbid solution and inhibit penetration of light thereby decreasing the rate of regeneration of Fe2+ and OH% radicals. As a result, the degradation rate will also be hampered at high initial PhACs concentration (Li et al., 2015; Mirzaei et al., 2017).

5.5.3.3. Influence of current density/intensity. Current density (J) and current intensity (I) are the controlling parameters in the EF process as it controls the generation rate of H2O2, cathodic generation of Fe2+ from consumed Fe3+. It also generates Cl− ions that have the oxidizing potential to degrade PhACs. Increasing J or I results in increased production of OH radicals, which lead to enhanced degradation of PhACs(Antonin et al., 2015; Garcia-Segura et al., 2014; Moreira et al., 2017). Although the increase in J or I increase the removal rate, very high J or I are not desirable (Moreira et al., 2017). Researchers have found that degradation of tetracycline and naproxen decreased when the value of J was above 70 mA/cm2 and 50 mA/cm2, respectively. That may be due to an increase in parasitic reactions, which leads to the formation of metabolites that consumed the available OH radicals (Coria et al., 2016; Liu et al., 2013).

5.5.3. Electro Fenton (EF) process In the EF process, in-situ H2O2 is electro generated, which is the main advantage over conventional and photo-Fenton processes (Ahmed et al., 2017). In-situ generation of H2O2 and reusability of the catalyst has made this process convenient by reducing the cost and eliminating the external addition of H2O2. It also enables only selective target reaction and overcomes the risks associated with transportation of H2O2( Wang et al., 2016a). In this process, the regeneration of Fe2+ from consumed Fe3+ occurs at the cathode, making it an environmentfriendly process (Moreira et al., 2017). Similar to degradation in anodic oxidation, degradation by electro-Fenton process occur in two ways (Brillas, 2014). Both M(OH)% on the anode and OH% radicals present in the bulk medium can degrade the PhACs (Sociedad Química de México., 2014) (Eq. (6), Eq. (7), Eq. (17), Eq. (18) and Eq. (19)).

O2(g ) + 2 H+ + 2e

H2 O2

At cathode

5.5.4. Photo-electro Fenton (PEF) process This process is the same as the electro Fenton process, but simultaneous photo-irradiation is carried out to regenerate consumed Fe2+ and OH% radicals that make this process more suitable than other Fenton processes (Ahmed et al., 2017; De Luna et al., 2012). PEF is mainly of two types (UV and solar based). Some modifications using ultrasound radiation has also been done recently (Babuponnusami and Muthukumar, 2014; Moreira et al., 2017). PEF process is more versatile and efficient than EF process as it overcomes the drawbacks like unintended consumption of OH radicals in parasitic reactions. It also uses the consumed Fe2+ for further oxidization of H2O2 as shown in Eq. (20), Eq. (21), and ) (Babuponnusami and Muthukumar, 2014; Moreira et al., 2017; Wang et al., 2016a).

(17) (18)

Fe3 + + e

Fe 2 +

At cathode

(19)

Other than the already mentioned influencing parameters for conventional Fenton process, EF process will also depend on current intensity/density, types of electrode material, types of electrolyte, the concentration of electrolyte, etc (Sociedad Química de México., 2014).

(20)

5.5.3.1. Influence of electrolyte. Electrolytes such as sodium chloride (NaCl), potassium chloride (KCl), sodium sulfate (Na2SO4), sodium perchlorate (NaClO4), sodium nitrate (NaNO3), and sodium carbonate (Na2CO3) are commonly used as supporting media for EF processes (Moreira et al., 2017). Some researchers found that NaCl was more efficient while others found that Na2SO4 was more efficient for EF processes. In the case of mineralization, Na2SO4 is superior to NaCl because, in the case of NaCl, chloro-derivatives are formed, which inhibits the degradation process (Ghoneim et al., 2011; Thiam et al., 2015, 2014). NaClO4 shows better suitability than NaCl because ClO4− will not form complexes with Fe2+/Fe3+ and will not react with OH% radicals (Daneshvar et al., 2008; De Laat et al., 2004). Very few studies

(21) (22) 5.5.4.1. Influence of photo-radiation and initial concentration. From the previous discussion as in photo Fenton process, it can be said that with an increase in radiation intensity higher degradation rate, mineralization rate and suitability near neutral pH have been possible (Aleksić et al., 2010; Wang et al., 2016b). Increase in initial concentration will produce a turbid solution that will check the penetration of light and hamper the degradation rate and regeneration of Fe2+ (Li et al., 2015; Mirzaei et al., 2017).

17

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Table 9 Influencing parameters and responsible radicals in degradation of selected PhACs by Fenton processes.

Table 10 Experimental conditions and removal efficiencies of PhACs in synthetic wastewater by Fenton based processes. AOPs/PhACs

Experimental conditions

Conventional Fenton Process Ciprofloxacin Dose: 5:1 nZVI:H2O2; pH: 7; Room temp; Time: 120 min; [Co]: 1 × 104 μg/l Trimethoprim Dose: 1:3 M ratioiron/oxalate; Fe3+ 5.0 mg/l and the same dose of H2O2; pH: 6.5; Time: 120 min; [Co]: 2 × 104 μg/l Sulfamethoxazole Dose: Fe2+ = 2.6 mg/l and H2O2 = 30 mg/l; pH: 2.52-.8; Temp: 20 °C; Time: 420 min; [Co]: 1 × 104 μg/l Tetracycline Dose: Fe2+ = 1 g/l and H2O2 = 150 mM; pH: 3.7; Temp: 22 °C; Time: 60 min; [Co]: 1 × 105 μg/l Paracetamol Dose: 6 g/L magnetite and 28 mM H2O2; pH: 2.6; Temp: 60 °C; Time: 300 min; [Co]: 1 × 105 μg/l Ibuprofen Dose: 1.2 mM Fe2+and 0.32 mM H2O2; pH: 3; Temp: 30 °C; Time: 120 min; [Co]: 1.8 × 105 μg/l Naproxen Diclofenac Ketoprofen Atenolol Metoprolol Caffeine Carbamazepine Estrone 17-β estradiol Photo Fenton Ciprofloxacin Trimethoprim Sulfamethoxazole Tetracycline Paracetamol Ibuprofen Naproxen Diclofenac Atenolol Metoprolol Caffeine

Removal Efficiency (%)

Reference

99.3 98

Mondal et al. (2018) Dias et al. (2014)

100

(Murillo-Sierra et al., 2018; Wang et al., 2011a, 2011b) (Liu et al.,2016a) Velichkova et al. (2013) (Klamerth et al., 2010; MéndezArriaga et al., 2010) Paiva et al. (2018)

100 100 100

Dose: 4.83 mg/l Fe2+and 9.98 mM H2O2; Ultrasound radiation; pH: 3; Temp: 283-3 °C; Time: 10min; [Co]: 2 × 104 μg/l Dose: 0.05 mM Fe2+, 200–400 mg/l H2O2; pH: 7; Temp: 20–40 °C, [Co]: 4.55 × 104 μg/l Dose: 0.1 mM Fe2+, 0.1 mM H2O2; pH: 3; Temp: 20 °C; Time: 5 min; [Co]: 1.27 × 104 μg/l Dose: 5 mg/l Fe2+ and 100 mg/l H2O2; pH: 7; Temp: 35 °C; Time: 150min; [Co]: 2 × 104 μg/l

100

Dose: 2.8 mg/l Fe2+ and 95 mg/l H2O2; pH: 7; Temp: 35 °C; Time: 150 min; [Co]: 2 × 104 μg/l Dose: 3 μmol/l Fe2+ and 10.0 μmol/l H2O2; pH: 3; Time: 30 min; [Co]: 1000 μg/l Dose:1.25 × 10−5 mol/L Fe2+ and 1.39 × 10−4 mol/L H2O2; pH: 3.52; Temp: 25 °C; Time: 120 min; [Co]: 4.98 × 103 μg/l Dose: 0.28 μM Fe3+ and 1.66 mM H2O2; pH: 3, Room temp; Time: 160 min; [Co]: 5 × 103 μg/l Dose: 1 mM Fe3+; pH: 3; Temp: 35 °C Time: 60 min; [Co]: 1 × 103 μg/l

100 95 100

Real et al. (2009) Real et al. (2009) (Isarain-Chávez et al., 2010; Veloutsou et al., 2014) Romero et al. (2016) Oliveira et al. (2015a, 2015b) Domínguez et al. (2012)

98.4 > 90

Feng et al. (2005) Ifelebuegu and Ezenwa (2011)

Dose: 5:1 nZVI:H2O2; UV 280 nm; 6W; pH: 7; Room temp; Time: 45 min; [Co]: 1 × 104 μg/l; Volume 500 mL Dose: 5 mg/l Fe2+ and 2.6 mM H2O2; UV lamp; pH: 2.8; Temp: 25 °C; Time: 60 min; [Co]: 2 × 104 μg/l Dose: H2O2=300 mg/l and Fe2+=10 mg/l; UV 365 nm; pH: 2.8; Temp: 25 °C; Time: 60 min; [Co]: 2 × 105 μg/l Dose: 0.2 mM (K3Fe(C2O4)3.3H2O; H2O2= 10 mM; Black light lamp 365 nm; 15W; pH: 2.5; Time: 60 min; [Co]: 2.4 × 104 μg/l; Volume 500 mL Dose: [Fe2+]=500 mg/l and [H2O2]: 1500 mg/l; pH: 3; UV 10W; Volume 4L; Time: 120 min; [Co]: 1 × 105 μg/l Dose: 5 mg/l Fe2+and 50 mg/l H2O2; pH: 7.6–8.0; Temp: 37 °C; Time: 90 min, [Co]: 100 μg/l; Xe lamp (290–350 nm) Dose: 1 mg/l Fe2+and 2 mg/l H2O2; Dark light lamp 350–400 nm; pH: 2.6; Temp: 252-8 °C; Time: 30 min; [Co]: 900 μg/l Dose: 15 mM H2O2 and 0.05 mM Fe2+; Solar irradiation; pH: 7; Temp: 30–40 °C; Time: 110 min; [Co]: 5 × 104 μg/l Dose: 2.8–5 mg/l Fe2+; 95–100 mg/l H2O2; Hg lamp 290 nm; pH: 2.9; Time: 180 min; [Co]: 2 × 104 μg/l Dose: 10 mg/l Fe2+ and 150 mg/l H2O2; Black light 365 nm; pH: 3; Temp: 25 °C; Time: 150 min; [Co]: 2 × 104 μg/l Dose: 10.0 mg/l Fe2+ and 42.0 mg/l H2O2; UV lamp; pH: 2.5; Time: 120 min; [Co]: 5.2 × 104 μg/l

100

Mondal et al. (2018)

100

Dias et al. (2014)

100

González et al. (2007)

100

Bautitz and Nogueira (2007)

100

Alalm et al. (2015)

100

100

(Klamerth et al., 2010; MéndezArriaga et al., 2010) (Agüera et al., 2005; Paiva et al., 2018) Pérez-Estrada et al. (2005)

100

Veloutsou et al. (2014)

100

(Romero et al., 2016; Veloutsou et al., 2014) Trovó et al. (2013)

100 100 100

100

98.54

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Table 10 (continued) AOPs/PhACs

Experimental conditions

Removal Efficiency (%)

Reference

Carbamazepine Estrone

Dose: 2:1 Persulfate:Fe2+; UV-Vis system; Time: 30 min; [Co]: 1.2 × 107 μg/l Dose: 20.8 μmol/l Fe3+ and 1664 μmol/l H2O2; Metal halide lamp 313 nm; 250W; pH: 3; Time: 160 min; [Co]: 5 × 103 μg/l; Volume 50 mL Dose: 5 g/l α-FeOOH and 9.7 mmol/l H2O2; pH: 7.47; UV 30W; Temp: 20 °C; Time: 480 min; [Co]: 272 μg/l; Volume 1L

100 98.4

Ahmed and Chiron (2014) Feng et al. (2005)

> 90

Yaping and Jiangyong (2008)

Carbon felt cathode and Pt anode; 0.1 mM Fe2+; Electrlyte: 0.05 M Na2SO4; Current intensity: 400 mA; pH: 3; Temp: 25 °C; Time: 40 min; [Co]: 5.3 × 104 μg/l Trimethoprim Dose: 2 mg/l Fe2+; Electrode; 7.0 g/l Na2SO4; pH: 3.5; Temp: 20 °C; Time: 60 min; Current density: 5 mA/cm2; [Co]: 2 × 104 μg/l Sulfamethoxazole Dose: 0.2 mM Fe2+ and Electrode; Current intensity:300 mA; Electrolyte: 0.05 M Na2SO4; pH: 3; Temp: 23 °C; [Co]: 5.2 × 105 μg/l Tetracycline BDD or Pt anode and carbon felt cathode; Electrolyte: 0.05 M Na2SO4; 2 g/l pyrite; Time: 40 min; Current intensity: 300 mA; pH: 3; [Co]: 9 × 104 μg/l Paracetamol 2 stainless steel cathodes and Ti coated RuO2/IrO2-coated DSA anode; Fe2+= 0.087 mM; Current density: 38 A/m2; pH: 3; Time: 120 min; [Co]: 7.6 × 105 μg/l; 48W; Volume 3.5L Ibuprofen BBD anode and O2 diffused cathode; 0.5 mM Fe2+; Electrolyte: 0.05 M Na2SO4; Current density: 33.3 mA/cm2; pH: 3; Time: 40 min; [Co]: 4.1 × 104 μg/l Naproxen BDD anode and air-diffusion cathode; 0.50 mM Fe2+; Electrolyte: 0.050 M NaClO4; Current density: 50 mA/cm2; pH: 3; Temp: 35 °C, [Co]: 4 × 104 μg/l Diclofenac Gas diffused cathode; 10 mM/l FeSO4; Current intensity: 4A; Time: 120 min; [Co]: 2 × 105 μg/l Ketoprofen Iron electrodes; 50 mM Fe2+, Current intensity: 300 mA; Electrolyte: 0.03M Na2SO4; Time: 60 min; pH: 3; [Co]: 2.5 × 103 μg/l Atenolol 4 electrode Pt & BDD anodes and Air diffused Carbon PTFE & carbon felt cathodes; 0.5 mM Fe2+; Electrolyte: 0.05M Na2SO4; Current intensity: 12 mA; pH: 3; Time: 30 min; [Co]: 1 × 105 μg/l Metoprolol Dose: 0.5 mM Fe2+ and 0.1 mM Cu2+; Current intensity: 120 mA; pH: 3; Temp: 35 °C; Time: 30 min; [Co]: 6.7 × 104 μg/l Caffeine BDD anode; Fe2+= 0.2 mM; Electrolyte: 0.05 M Na2SO4; Current intensity: 300 mA; pH: 3; Temp: 21 °C; Time: 10 min; [Co]: 2 × 104 μg/l Carbamazepine BDD anode; 0.25 mM Fe2+; Electrolyte: 1 g/l Na2SO4; pH: 3; Time: 120 min; [Co]: 1.2 × 104 μg/l 17-β estradiol Dose: Pt anode, 0.2 mM Fe2+; Electrolyte: 0.05 M Na2SO4; Current intensity: 200 mA; pH: 3; Time: 70 min; [Co]: 1 × 103 μg/l Photo electro Fenton Ciprofloxacin Microwave discharge electrode less lamp; 1.0 mM Fe2+; Electrolyte: 0.05 M Na2SO4; Current density: 18 mA/cm2; pH: 3; Time: 45 min; [Co]: 2 × 105 μg/l; Erythromycin Pt/Ti plate anode and graphite-felt cathode; 0.50 mol/l Fe2+; Electrolyte: 0.050 mol/l Na2SO4; Solar light; Current density: : (−)0.16 mA/cm2; pH: 3; Time: 300 min; [Co]:1.6 × 105 μg/l Trimethoprim BDD anode and carbon–PTFE air-diffusion cathode; Fe3+ = 2 mg/l; Electrolyte: 50 mM Na2SO4; pH: 4.5; Temp: 20 °C; Time: 120 min; Current density: 5 mA/cm2; [Co]: 2 × 104 μg/l, UVA irradiation Sulfamethoxazole BDD anode, carbon-PTFE air diffusion cathode; Electrolyte: 5% v/v H2SO4 in 0.05 M Na2SO4; Current intensity: 47 mA/cm2; pH: 3; Time: 120 min; [Co]: 5 × 104 μg/l, Solar irradiation Tetracycline Fe3O4–graphite cathode; 500W; Electrolyte: 10 g/l Na2SO4; Current density: 70 mA/cm2; pH: 3; Time: 150 min; [Co]: 20 μg/l; Volume 100 mL, UV lamp 360 nm Ibuprofen BBD anode and O2 diffused cathode; 0.5 mM Fe2+; Electrolyte: 0.05 M Na2SO4; Solar radiation; current density: 6.6mA/cm2; pH: 3; Time: 40 min; [Co]: 4.1 × 104 μg/l Naproxen BDD anode air-diffusion cathode; 0.50 mM Fe2+; Electrolyte: 0.050 M NaClO4; UV 360 nm; Current density: 50 mA/cm2; pH: 3; Temp: 35 °C, [Co]: 4 × 104 μg/l Atenolol 4 electrode Pt & BDD anodes and Air diffused Carbon PTFE & carbon felt cathodes; 0.5 mM Fe2+; Electrolyte: 0.05M Na2SO4; UVA 365 nm; Current intensity: 12 mA; pH: 3; Time: 30 min; [Co]: 1.58 × 105 μg/l Metoprolol BDD/ADE; UVA; 0.5 mM Fe2+; Current intensity: 120 mA; pH: 3; Time: 30 min; [Co]: 6.7 × 104 μg/l

100

Yahya et al. (2014)

100

Moreira et al. (2014)

100

( Dirany et al., 2010)

100

(; Liu et al., Liu et al., 2016.)

98

De Luna et al. (2012)

100

Skoumal et al. (2009)

100

Coria et al. (2016)

99.4 100

Rocha et al. (2009) Alizadeh Fard and Barkdoll (2018)

90

Isarain-Chávez et al. (2010)

100

Isarain-Chávez et al. (2011)

100

Ganzenko et al. (2015)

73 100

Komtchou et al. (2015) Naimi and Bellakhal (2012)

80

Wang et al. (2018a, 2018b)

78

Pérez et al. (2017)

100

Moreira et al. (2015)

100 100

(Murillo-Sierra et al., 2018; Wang et al., 2011a, 2011b) (Liu et al., 2013)

100

Skoumal et al. (2009)

100

Coria et al. (2016)

97

Isarain-Chávez et al. (2010)

100

Isarain-Chávez et al. (2011)

17-β estradiol Electro Fenton Ciprofloxacin

6. Physical processes

et al., 2012; Kwon et al., 2012; Shao et al., 2018). The process is characterized by high degradation efficiency and feasible electrical energy. However, He et al. (2014) noted that although diclofenac concentration was brought down below detection limit, there was no significant reduction in TOC(He et al., 2014). The main influencing factor in this process is the irradiation dose (kGy) which are listed in Table 11. The high establishment cost of the system for the process of water treatment is also a downside to this process.

6.1. Electron beam Radiation processes like usage of electron beams and gamma rays for degradation of organic compounds have been studied since the 1980s (Miklos et al., 2018). The accelerated electrons from an electron beam source (0.01meV-10meV) penetrates the water surface and leads to the formation of hydroxyl radicals, hydrogen radicals (H%) and hydrated electrons (e-eq) which helps in degrading the organic compounds. The penetration depth of the accelerated electrons are directly proportional to the energy of the incident beam. PhACs like sulfamethoxazole, tetracycline, diclofenac, iopromide, and other antidepressants have been efficiently degraded by electron beams (He et al., 2014; Kim

6.2. Microwave discharge electrodeless lamp Microwaves are highly energetic radiations in the range of 300 MHz to 300 GHz. Their application in degrading organic contaminants from water has been examined in recent times (Miklos et al., 2018). 19

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However, only a few studies have been conducted in the field of degradation of PhACs. Microwave discharge electrodeless lamps have been used to generate UV-C radiations. As a result, they have been used in combination with Fenton process to degrade ciprofloxacin ( Wang et al., 2018a). The microwave/photo-Fenton provided around 98% ciprofloxacin reduction and around 90% TOC reduction in 2h and 6h, respectively. The major drawback to this process is that the microwave radiations are converted to heat and a cooling equipment is required to prevent the treated water from getting overheated.

Table 11 Experimental conditions and removal efficiencies of PhACs in synthetic wastewater by physical oxidation process. AOPs/PhACs

Experimental Condition

Removal Efficiency (%)

References

Dose 19.6 kgY; 2.5 meV; 100 kW Dose 0.5 kgY; 10 meV; 10 mA; 100 kW Dose 1 kgY; 1 meV; 40 kW Dose 0.6 kgY; 1 meV; 40 kW Dose 2 kgY; 1.8 meV; 0–10 mA

99 99

Kwon et al. (2012) He et al. (2014)

88.6

Kim et al. (2012)

99

Kim et al. (2012)

99

Shao et al. (2018)

30W 30W 30W 30W 30W 30W 30W 30W 30W 30W 30W 30W 30W 30W 60min

19 99 93 99 99 50 99 99 99 89 99 99 91 99 100

Estradiol

60min

100

Ibuprofen

60min

80

Carbamazepine

60min

80

trimethoprim

60min

80

diazepam

60min

50

Diclofenac Carbamazepine Sulfamethaxazole Ultrasound Ciprofloxacin

40min 40min 40min

71 67 54

Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Ajo et al. (2018) Banaschik et al. (2018) Banaschik et al. (2018) Banaschik et al. (2018) Banaschik et al. (2018) Banaschik et al. (2018) Banaschik et al. (2018) Back et al. (2018) Back et al. (2018) Back et al. (2018)

0.2 kW; 544 kHz; volume 1L; 120min 0.8 kW; 45 kHz; volume 50 mL; 180min 0.4 kW; 45 kHz; volume 1L; 60min 0.4 kW; 45 kHz; volume 1L; 60min 0.8 kW; 45 kHz volume 50 mL; 180min 0.12 kW; 577 kHz; volume 250 mL; 20min 0.12 kW; 861 kHz; Volume 500 mL; 30 min

60

Electron Beam Iopromide Diclofenac Sulfamethaxazole

6.3. Plasma discharge

Cholrotetracyline

Non-thermal plasma (NTP) involves ozone formation from ambient oxygen and hydroxyl radical production from water molecules. Strong electric fields are applied between water and gas phase which initiate both chemical and physical processes (Miklos et al., 2018). Other than the oxidizing radicals, shock waves and UV radiation are evolved, which helps in the degradation process. The gas phase pulsed corona discharge method is the most energy efficient NTP for water treatment. The free gaseous electrons and the non-equilibrium state of the ions are responsible for negligible heat losses in the process and thus making it highly energy efficient (Ajo et al., 2018). Ajo et al. (2018) evaluated the efficiency of a gas-phase pulsed corona discharge treatment on hospital effluent of Finland. At 30W power and frequency of 840Hz, most of the PhACs got completely degraded, except ibuprofen, and caffeine which were removed by 50% and 19%, respectively. The decreased removal for ibuprofen and caffeine may be because of their molecular structure preventing oxidation (Ajo et al., 2018). In another study, high voltage pulses of 80 kV with a frequency of 20Hz was used to generate corona plasma to completely degrade of diclofenac and estradiol in 1h. However, ibuprofen, carbamazepine, trimethoprim showed 80% removal and diazepam showed 50% removal after 1h of treatment (Banaschik et al., 2018). At a fixed voltage of 16 kV and frequency of 500Hz the removal efficiencies of diclofenac, carbamazepine and sulfamethoxazole were 34%, 29%, and 25%, respectively. When the frequency was increased to 2000Hz the removal efficiencies of the compounds increased to 71%, 67%, and 54%, respectively. This shows frequency had a significant impact on degradation of the PhACs(Back et al., 2018).

Floxentine Plasma Discharge Caffeine Carbamazepine Ciprofloxacin Estriol Estrone Ibuprofen Metoprolol Naproxen Ofloxacin Paracetamol Propranolol Sulfamethoxazole Tetracycline Trimethoprim Diclofenac

6.4. Ultrasound

Triclosan

Ultrasound of frequency (20–500 kHz) is used to sonicate water and induces compression and rarefaction leading to the formation and collapse of micro-bubbles. When these micro-bubbles collapse they generate high temperature, pressure, and highly reactive radicals like hydrogen and hydroxyl radicals. These radicals react with the organic compounds to degrade them. This process is almost independent of the water matrix, as a result, the presence of other ions in the wastewater will not affect the degradation. At 544 kHz of ultrasound frequency around 40% degradation of ciprofloxacin was achieved in 2h (De Bel et al., 2011). Naddeo et al. (2013) treated 15 PhACs with sonication. The average final concentration of the individual PhACs was around 70% after 180 min of sonication at 45 kHz (Naddeo et al., 2013). In another study, almost complete degradation of ibuprofen was observed after sonication for 20 min at 577 kHz frequency. However, removal of TOC was only limited to around 10% suggesting complete breakdown was not achieved in this process. The effect of initial concentration of PhACs was observed when a 15 μM M solution of diclofenac was completely degraded in 0.5 h at 861 kHz, but when the concentration of diclofenac was increased to 70 μM and 130 μM degradation was limited to only about 50% under the same experimental conditions (Güyer and Ince, 2011). The degradation of PhACs can significantly improve if this process is combined with other oxidation processes like photocatalysis, ozone, and Fenton processes (Kanakaraju et al., 2018).

carbamazepine Diclofenac Erythromycin Ibuprofen Diclofenac

95 50 50 50 99 99

De Bel et al. (2011) Naddeo et al. (2013) Naddeo et al. (2013) Naddeo et al. (2013) Naddeo et al. (2013) Ziylan-Yavas and Ince (2018) Güyer and Ince (2011)

7. Transformation products and their toxicity PhACs undergo various degree of transformation during treatment by various AOPs. The parent compound is only partially degraded which reduces its toxicity and increases its biodegradability. However, it may be transformed into another product which are highly persistent and inert in nature. These compounds may be toxic in nature and have more detrimental effects than the parent compound. When they enter the food chain, they pose additional ecological risks. The PhACs undergo various abiotic transformation processes during AOPs (e.g. photochemical degradation, photo-excited organic matter, reaction with 20

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singlet oxygen, hydroxyl radicals, sulfate radicals, and other reactive species (Sharma et al., 2018). Erythromycin-H2O is a transformation product of the antibiotic erythromycin and have been detected in many WWTP effluent of major countries at concentrations around 6 μg/L (Evgenidou et al., 2015). Hydrolysis products of other antibiotics like amoxicillin (two diastereomers of amoxilloic acid and two diastereomers of amoxicillin diketopiperazine-2′,5′), trimethoprim ((2,4-diaminopyrimidin-5-yl) (3,4,5trimethoxyphenyl)methanone and 2,4-diaminopyrimidine-5-carbaldehyde) were also found in WWTP effluent of Spain and Greece, respectively (Evgenidou et al., 2015). During photolysis and photocatalysis, ciprofloxacin may undergo decarboxylation, defluorination, and loss of the piperazine moiety leading to 5 transformation products which were not biodegradable (Fatta-Kassinos et al., 2011). Upon treatment with UV light at 355 nm, tetracycline got degraded to form 2 major intermediate products having m/z value of 398 and 413.9. These intermediates had the naphthol ring on tetracycline intact and was found to be toxic (Jiao et al., 2008). The photochemical transformation of sulfamethoxazole was characterized by the splitting of the sulfonamide bond and the rearrangement of isoxazole ring. These transformation products were found to have higher toxicity than sulfamethoxazole (Fatta-Kassinos et al., 2011). Upon photocatalytic treatment, trimethoprim underwent demethylation, hydroxylation, and splitting of original compound and the formed products were of moderately toxic nature (Sirtori et al., 2010). Among analgesics, ibuprofen had 3 transformation products (1-hydroxy-ibuprofen (1′–OH–IBU), 2-hydroxy-ibuprofen (2′–OH–IBU), carboxy-ibuprofen (CX-IBU)), and diclofenac had 3 transformation products (5-hydroxy-diclofenac (5′–OH–DCF), 4-hydroxy-diclofenac (4′–OH–DCF), 4-hydroxy-diclofenac-dehydrate (4′–OH–DCF-H2O)), which were found in influent and effluent (Evgenidou et al., 2015). In presence of H2O2 hydroxylated-ibuprofen byproducts are formed. During photo-Fenton process ibuprofen undergoes decarboxylation and complexes are formed with iron. When diclofenac was degraded using photo-Fenton, ozonation and catalytic oxidation in presence of H2O2 quinone and phenol compounds were formed (Kosjek and Heath, 2008). The intermediate products of diclofenac formed during photolysis were found to be phytotoxic in nature (Fatta-Kassinos et al., 2011). Hydroxyl radicals react with paracetamol to form hydroquinone. The OH radical gets attached to the ispo position of the phenyl ring of paracetamol and the –NHCOCH3 bond gets split (Kosjek and Heath, 2008). Upon Fenton oxidation of paracetamol, a transformation product with m/z value of 198 was formed and it posed high toxicological risk (Cuervo Lumbaque et al., 2018). The photo-transformation of naproxen lead to the formation of dimeric compounds and they were found to be more toxic than the parent compound (Fatta-Kassinos et al., 2011). One major by-product of oxidation of carbamazepine is carbamazepine-10, 11 epoxide which is an active compound with anticonvulsant properties (Evgenidou et al., 2015). Another transformation product of carbamazepine which is found in WWTP effluent is iminostilbene which is genotoxic in nature. Upon O3 treatment, carbamazepine underwent an increase in mass-to-charge values by 14, 30 and 46. The transformed product exhibiting the m/z value of 266 was identified to be 1-(2-benzaldehyde)-(1H, 3H)-quinazoline-2,4-dione. Acridine, another genotoxic transformation product of carbamazepine was formed as a result of photolysis and TiO2 photocatalysis (Kosjek and Heath, 2008). During reaction with hydroxyl radicals, the β-blockers undergo hydroxylation. There is an addition of an OH bond in the ispo position of the aromatic group and the ether bond gets split (Kosjek and Heath, 2008). Metabolites of metoprolol (a-hydroxymetoprolol (a-HMTP) and O-desmethylmetoprolol (O-DMTP)) were also detected in WWTP effluent of Spain (Rubirola et al., 2014). There were 3 major transformation products formed for propranolol after Fenton processes having m/z value of 134, 292 and 266. The transformation products having m/ z values of 134 and 292 were persistent compounds while the later

posing high toxicological risk (Cuervo Lumbaque et al., 2018). During treatment of estrogens by any AOP involving hydroxyl radicals, the degradation can follow two pathways. The hydroxyl radicals may attack the aromatic ring or the aliphatic ring. During degradation of 17-β estradiol if the hydroxyl radicals attack the aliphatic ring it may give rise to intermediates having ketones, alcohol and olefins groups. The probable transformation products can be estriol, estra-1,3,5 (10)triene-3,6α,17β-triol and (17β)-estra-1,3,5(10),9(11)-tetraene-3,17diol. When the aromatic compounds are attacked dihydroxy and quinone like products are formed. These compounds further go ring rupturing reactions and are attached by more hydroxyl radicals and can get completely mineralized (Yaping and Jiangyong, 2008). In another study, 17-β estradiol was transformed into quinone methide. Ozonation of 17-β estradiol led to the formation of 10ε-17β-Dihydroxy-1,4-estradieno-3-one and 2-hydroxy-estradiol while activated sludge process converted it into estrone (Bila et al., 2007; Skotnicka-Pitak et al., 2008). 8. Comparison of advanced oxidation processes 8.1. Removal efficiency and time PhAC degradation by AOPs depend on a number of factors including method of activation, presence of catalysts, intensity or dose of irradiation or electricity, etc. As a result, different AOPs have varying treatment time and removal efficiency. The removal efficiency and time of treatment pertaining to different AOPs have been illustrated in Fig. 3. Degradation by ozonation is one of the quickest among all the AOPs with an average degradation time of 1h. Since degradation largely depends on the ozone doze applied the maximum time for degradation was found to be 4h and the minimum time for degradation was as low as 10 min. The average removal efficiency for this process was around 90%, making this a very good treatment option for implementing in WWTPs. The removal efficiency and time for photocatalysis depends on a number of parameters, which includes the type of light used, type of catalyst used, dose of the catalyst, etc. PhACs have shown efficient degradation for all the AOPs. Most of the PhACs got degraded within 4h during photocatalytic treatment (An et al., 2010; Barbara Ambrosetti et al., 2015; Molinari et al., 2017; Zhu et al., 2016; Xekoukoulotakis

Fig. 3. Overview of treatment time and removal efficiencies of PhACs by different AOPs. Data from: Table 4, Table 8, Tables 10 and 11. 21

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et al., 2011). Antibiotics, antiepileptic drugs, and β-blockers were degraded most efficiently having an average removal efficiency of 90% (An et al., 2010; Barbara Ambrosetti et al., 2015; Molinari et al., 2017; Xekoukoulotakis et al., 2011; Molinari et al., 2017, 2017; Chen et al., 2018a, 2018b;He et al., 2017b). High removal of hormones and stimulants could not be achieved (Arfanis et al., 2017; Czech and Rubinowska, 2013). Time taken to degrade PhACs by anodic oxidation was comparatively less as compared to photocatalysis (Fig. 3). Escudero et al. (2017) have done a comparative study between photocatalysis and anodic oxidation to remove the same pollutant and the rate constants obtained were 2.75 × 102−min−1 and 3.45 × 103− min−1 respectively, further establishing the fact that anodic oxidation is a faster process (Escudero et al., 2017). The reason may be because M(OH%) radicals generated in anodic oxidation are more reactive than OH radicals generated in photocatalysis. The degradation of antibiotics and β-blockers took more time, which may be because of their high pKa values. At neutral or acidic pH they tend to be positively charged and hence they do not easily diffuse and come in contact with the positive anode. Analgesics had the maximum removal efficiency and antiepileptics had the least removal efficiency. During oxidation by activation of PS/PMS, the average time taken (2–3.5h) for degradation was the lower than photocatalysis and anodic oxidation (Fig. 3). This was because SO4% involved in this process have the highest oxidizing potential values among the other radicals responsible for degradation in other processes. Also, in this process along with SO4%, OH radicals act simultaneously in degrading the contaminants. The removal efficiency observed for antibiotics, analgesics and stimulants were better than photosynthesis and anodic oxidation. However, antiepileptics, hormones, and β-blockers showed reduced removal efficiency. Although during conventional Fenton process most PhACs get degraded in the early stage of the process, complete removal cannot be ensured without complete mineralization (Alalm et al., 2015; Li et al., 2015). It was observed that almost all the PhACs took at least 2–3h for degradation except tetracycline, diclofenac, and 17-β estradiol, which got removed in less than 1h. This may be accounted for simultaneous adsorption and degradation (Bae et al., 2013; Hou et al., 2016; Ifelebuegu and Ezenwa, 2011). From Fig. 3, it can be inferred that average reaction time taken for the complete removal of PhACs in conventional Fenton process lies between 0.5h and 3h. However, paracetamol took about 5h to achieve complete degradation (Trovó et al., 2009; Velichkova et al., 2013). Photo Fenton processes showed less average degradation time (1–2.5h) than conventional Fenton processes because of the presence of more OH radicals. All the PhACs get readily degraded except β-blockers and 17-β estradiol that took around more time for complete degradation. The higher time (2.5–3h) observed by Veloutsou et al. (2014) may be due to consumption of OH radicals by other organic and inorganic complexes formed at intermediate stages (Veloutsou et al., 2014). 17-β estradiol took around 8h to achieve around 90% removal. This could be because the solution pH was kept at 7.47 for the photo Fenton process but as discussed earlier this process works best in acidic pH (Yaping and Jiangyong, 2008). EF processes showed improved removal efficiency and better average reaction time (< 2h) than conventional and photo Fenton processes. However, carbamazepine showed only around 73% removal (Komtchou et al., 2015). Since the solution pH was maintained at 3, and pKa value of carbamazepine is 13.9, it may acquire a net positive charge. It may be due to the repulsion of like charges that carbamazepine was not able to come in proximity with the positively charged anode surface. As a result, complete degradation was not achieved (Komtchou et al., 2015; Moreira et al., 2017). Caffeine also took more time to get completely degraded (5–6h). Since stimulants exist as positively charged particles at acidic pH, they do not get readily attacked by electrophilic OH radicals (Ganzenko et al., 2015; Murugananthan

Fig. 4. Overview of EEO values of different AOPs during PhAC degradation. Data from: Table 8, Tables 10 and 11.

et al., 2007; Santos et al., 2010). As a result, degradation took more time. PEF process requires less reaction time than other Fenton process for complete removal. It was found that the PEF process shows the almost complete removal of PhACs. However, ciprofloxacin and erythromycin demonstrated lower degradation (80% and 78%, respectively) which may be due to the consumption of OH radicals by the produced inorganic ions such as NH4+ and NO3− (Pérez et al., 2017; Wang et al., 2018a; Yahya et al., 2014). Beta-blockers require less reaction time (< 1h) than other PhACs which may be due to its high hydrophobicity and less solubility (Veloutsou et al., 2014). The time taken for degradation by the physical process were much lesser. Oxidation by electron beam took the minimum time of all the AOPs. Degradation of PhACs depend directly on the dose of irradiation dose (kGy) and an approximate time taken for 10 kGy irradiation dose in 3 min. High degradation of the PhACs was also achieved by this process, but complete mineralization was not observed. The removal efficiency using plasma discharge was high and the average time taken for degradation is also less than 1h. Degradation time by pure ultrasound was also comparatively less with an average time of 1.5h. However, the average removal efficiency was around 70% varying significantly from 50% to complete removal. 8.2. Electric energy per order (EEO) Energy efficiency of the AOPs is usually represented in the form of electrical energy per order. The main operating cost of the AOPs comes from running of UV lamps, electric current in anodes, and other activation processes involving electricity. Processes like conventional Fenton process and process not requiring any major electricity for operation other than stirring have not been considered in this study. Electrical energy per order is said to be the electrical energy required in kilowatt-hour (kWh) to reduce the concentration of the PhACs by 1st order of magnitude in 1 m3 of contaminated water (Asaithambi et al., 2017). The EEO (kWh/m3/order) was calculated using Eq. (23) and are shown in Fig. 4.

EEO =

22

P V

60

t log

(

1000 1

1 % removal

)

(23)

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where, P is the Power (kW), t is the treatment time (min), V is the volume of the contaminated water (L) and %removal (decimals) which are provided in Table S4. AOPs with 25th-75th percentile EEO values less than 10 kWh/m3/ order were plasma discharge and electron beam. With a high removal efficiency and low EEO values, these may have prospect in water treatment in WWTPs, but the drawbacks are high capital cost, maintenance cost and risk from potential risk from X-rays. Fenton processes and PS/PMS activation processes have 25th-75th percentile EEO values around 300 kWh/m3/order which are more power consuming, but do not have high capital cost. Photocatalysis and anodic oxidation processes have EEO values ranging from 10 kWh/m3/order to 13000 kWh/ m3/order. Upon proper optimization of the power of the light source and electricity by changing other process parameters, the values of EEO can be significantly brought down and can be up-scaled. Ultrasound alone consumes a lot of energy because of the high-frequency sound required for these processes. However, if integrated with other AOPs, the EEO values may be brought down and will also improve the performance of the integrated process. Ozone-based AOPs have been reported to have EEO values less than 1 kWh/m3/order (Miklos et al., 2018). The low EEO values and high removal efficiencies make this an attractive option for up-scaling.

•

•

•

9. Summary of findings The increasing health problems arising across the world has led to an increased utilization of pharmaceuticals, which has resulted in continual persistence of PhACs s in the aquatic ecosystem. The concentrations of many PhACs have already crossed the DWEL and poses an immediate threat to all kinds of aquatic life forms and human beings exposed to the contamination. The concentration of PhACs in water will keep on increasing with time since medicines will remain an indispensable component in the society. Many of the conventional municipal and sewage WWTPs are not equipped with technologies to bring down the concentration of PhACs below the DWEL. As a result, the performance of AOPs in degrading such pollutants have been investigated. Following points dictate the summary in brief.

•

• Ciprofloxacin was the highest occurring PhACs to be detected in

•

•

both wastewater and surface waters of Asia and their detected concentrations were much higher than the corresponding calculated DWEL. The concentrations of Atenolol and carbamazepine in WWTP influent of Asia and Europe respectively exceeded their corresponding DWEL. The detected concentrations of estrone in wastewaters of Europe and Africa were comparatively lower than most of the other occurring PhACs, but it crossed the DWEL because of its highly toxic nature. Diclofenac was found in both wastewater and drinking water in almost all the continents with concentrations greater than its DWEL. Furthermore, the PNECs of most of the PhACs in both wastewater and surface water were alarmingly lower than the reported concentrations. Hydroxyl radicals were the main driving radicals behind the degradation of most of the PhACs while h+ and O2% played second fiddle. However, for analgesics, h+ and O2% were found to have more significance in degradation than OH radicals. Since redox potential of SO4% is higher than all the other responsible radicals, degradation is enhanced upon activation of PS and PMS. HOCl% formed in anodic oxidation when the electrolyte is NaCl also helps in the degradation process. High removal of almost all PhACs (> 90%) was attained for all the AOPs. However, removal of stimulant by photocatalysis, analgesics by conventional Fenton Process and antiepileptics by almost all the processes were much less. Time taken by PS/PMS activation and Fenton-based processes are the least (2–3h). This is because of the higher redox potential of SO4% and continual generation of OH radicals in case of PS/PMS activation and Fenton-based processes,

respectively. Anodic oxidation took less time than photocatalysis because of generation of M(OH%) radicals. During photocatalysis, hormones, antiepileptics, and β-blockers have shown better removal at neutral pH while acidic pH favored degradation of antibiotics and analgesics. However, stimulants were degraded better at alkaline pH. Degradation by anodic oxidation was favored at near neutral pH. Alkaline pH showed maximum degradation for β-blockers, antibiotics, and analgesics when SO4% were used, but stimulants and antiepileptics were removed best at slightly acidic pH. Although degradation was maximum in acidic pH for all Fenton based processes, stimulants, hormones, and antiepileptics have shown efficient removal in near neutral pH for photo-Fenton process only. Various ions present in the water significantly affects the rate of degradation of AOPs. Organic matters and HCO3- ions were found to hinder the rate of degradation. Chlorine ions have also shown a negative effect on degradation in most cases except in case of degradation of analgesics and antibiotics by photocatalysis, β-blockers, and hormones by anodic oxidation. Presence of phosphate ions slows down degradation in case of photocatalysis while a slight enhancement in degradation was observed for β-blockers by PM/ PMS activation. Most of the PhACs do not get completely degraded after treatment with AOPs and are reduced to less toxic compounds. However, the transformation products of antibiotics like tetracycline, trimethoprim, and sulfamethoxazole were found to be toxic. Diclofenac and paracetamol also produced phytotoxic substances. Iminostilbene and acridine, two transformation products of carbamazepine were found to be genotoxic. Phototransformation of naproxen led to the formation of dimeric compounds, which were reported to be more toxic than the parent compound. Ozone-based AOPs have shown great potential in removing PhACs and the less energy required for their application makes it a good choice for upscaling. Photocatalysis, anodic oxidation, PS/PMS activation, and Fenton processes need more real water-based studies and process optimization to lower the operation cost. Physical processes like plasma discharge and electron beam are fast, efficient and operation cost is also less. However, the initial capital cost and risks of radiations make it unfavorable.

10. Future scope Most of the reported studies were conducted using synthetic wastewater and the influence of the presence of other contaminants (PhACs or other organic contaminants) have not been investigated in details. In synthetic wastewater, the initial concentration of PhACs considered (mg/L) was higher than occurring concentration in real wastewater (μg/L). Only a few research papers have shown the application of these processes in treating PhACs from hospital effluent or WWTP influent. A proper guideline is required to regulate the treatment processes of wastewater and the quality of drinking water. Degradation products formed should be analyzed and their toxicity should also be assessed because degraded products can be more toxic than the parent compound and complete degradation should be targeted. Further, degradation studies should be conducted on synthetic samples with mixture of various compounds for better understanding of the degradation mechanisms. Also, the performance of these synthetic water experimental studies should be evaluated with their performance in real wastewater. The high processing and maintenance cost of these AOPs should also be brought down to make it suitable for installing in the wastewater treatment plants. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.envres.2019.108542. 23

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