The Fate and Occurrence of Antibiotic-Resistant Bacteria and Antibiotic Resistance Genes during Advanced Wastewater Treatment and Disinfection: A Review

Abstract

Antimicrobial resistance (AMR) is a serious problem for modern society, not only associated with clinical environments, but also the natural environment. Conventional wastewater treatment plants (WWTPs) are important nodes for the dissemination of antibiotic resistance to the aquatic environment since they are reservoirs of antibiotic-resistant bacteria (ARB), antibiotic resistance genes (ARGs), and antibiotic residues. WWTPs are not designed to remove these antibiotic resistance determinants from wastewater, and as a result, they are present in treated effluent, leading to environmental and public health concerns regarding wastewater disposal and reuse. Additional treatments combined with conventional WWTPs can be barriers to the spread of AMR to the environment. In order to understand the effect of wastewater treatment methods on the removal of ARB and ARGs, an extensive bibliographic study was conducted. This review summarizes the efficiency of conventional disinfection methods, tertiary wastewater treatment, and advanced oxidation processes (AOPs) to remove ARB and ARGs from wastewater. In the context of the revised Urban Wastewater Treatment Directive 91/271/EEC, further studies are needed on the removal potential of AOPs on a full-scale, as they offer great potential for the removal of ARB and ARGs with a low formation of toxic by-products compared to conventional disinfection methods.

1. Introduction

Antimicrobial resistance is a crucial social and economic problem that is estimated to be responsible for 25,000 deaths per year in the European Union and 700,000 worldwide. Without any measures to hinder the alarming spread of antimicrobial resistance, it is estimated that 10 million people will die every year worldwide by 2050, exceeding the deaths due to the cancer [1]. The high consumption and misuse of antibiotics results in their continuous spread into the environment. The antibiotics used are not always fully metabolized by the body and are mostly excreted in their original form through urine and feces in wastewater treatment plants (WWTPs) [2]. Health care services are also a major contributor of antibiotics to municipal WWTPs [3]. The presence of antibiotic residues, even in subinhibitory concentrations, combined with the complex microbial communities (including pathogens and common bacteria) and nutrients facilitates the development of antibiotic-resistant bacteria and the promotion of the horizontal transfer of resistant genes [4]. The mechanisms of horizontal gene transfer (HGT) are conjugation, transformation, and transduction mediated by mobile genetic elements (MGE), such as plasmids, transposons, and integrons [5,6]. It was found that the conjugative transfer is enhanced considerably in the presence of a high quantity of nutrients [7]. Wastewater treatment plants are designed to remove nutrients and organic materials and, therefore, do not effectively remove antibiotics or bacteria and genes that carry antibiotic resistance [8]. The low concentration of antibiotics (ng/L–μg/L) in the bioreactors causes the selection of ARB and ARGs and favors the transfer of ARGs even to non-resistant bacteria grown during biological wastewater treatment [9,10,11]. Zhang et al. [12] and Mao et al. [13] studied the prevalence and proliferation of ARGs in full-scale activated sludge WWTPs and they noted that ARGs increased considerably in the aeration tanks due to the significant growth of bacteria. As a result, bioreactors could be hotspots for the propagation of antibiotic resistance in bacteria and reservoirs of ARB and ARGs [4,10,11]. In addition, some studies have found that biological treatment does not significantly reduce the relative abundance of ARGs [14,15]. Moreover, the primary treatment removes only slightly ARGs [14,16] while secondary settling tank removes higher fraction of ARGs, revealing that ARGs are transferred to sludge phase [15]. Apart from antibiotics, other chemical stressors, such as heavy metals and microplastics, are ubiquitous in the WWTPs, and these compounds could enhance the antibiotic resistance between bacteria [17,18,19,20,21]. Some researchers observed a high correlation between the sulII and czcA genes (cadmium, cobalt, and zinc resistance) in three WWTPs, resulting in the co-selection of ARGs and heavy metal resistant genes [22,23]. Most importantly, wastewater treatment plants are the main contributors to microplastics in the environment, as they collect and treat waste containing microplastic particles from personal care products and fibers from synthetic textiles that are released when clothes are washed [24]. ARGs, ARB, and pathogens could be adsorbed by microplastics, forming complex biofilms, and therefore enhancing the HGT between organisms [25,26,27].

Τhe treated effluents are discharged into surface waters, and therefore, wastewater treatment plants can be a significant barrier to the spread of antibiotic-resistant bacteria and genes [28] in the aquatic environment and subsequently in humans. Numerous studies reported the presence of ARB, ARGs, and facultative pathogenic bacteria in surface water [29,30,31]. It is vital to detect and hamper the spread of ARB and ARGs in the aquatic environment to protect public health, as it is directly related to wastewater reclamation, bathing waters, aquatic sports and the consumption of drinking water, fish, and irrigated plants [31].

The reuse of treated wastewater is expected to be an important aspect in mitigating water scarcity, especially in arid and semi-arid regions [32]. Regarding the European Union (EU), the potential for the reuse of wastewater is estimated to be six times greater than the applied one [33]. Although reclaimed water has higher quality requirements than wastewater discharged to aquatic receivers, the phenomenon of antimicrobial resistance is not taken into account at all. The regulation (EU) 2020/741 of the European Union for the reuse of wastewater is based on microbial indicators, and specifically, the maximum load of Escherichia coli in treated wastewater water intended for agricultural edible crops is 10 CFU/100 mL. This regulation does not take into account the load of ARB and ARGs, and thus, their presence is not monitored during the application of treated wastewater water to the soil and crops. Kampouris et al. [34] noted the presence of sulI, qnrS, tetM, and bla_OXA-58 genes in the soil was due to irrigation with secondary treated wastewater, whereas these genes were not present in the non-irrigated soil. Therefore, the irrigation with treated wastewater is a potential cause of the spread of ARB and ARGs in soil and crops via horizontal transfer [34,35]. Even at low concentrations, ARB and ARGs can cause detrimental effects in humans and animals. There is the possibility that they may be transferred through horizontal gene transfer, up the food chain, and, from there, to humans. In addition, some of the ARB and ARGs are pathogenic to humans, endangering their health [36].

There are numerous studies that have investigated the removal of ARB and ARGs with wastewater treatment processes. The objective of this review is to present the efficiency of tertiary wastewater treatment technologies (membrane filtration, granular activated carbon, sand filtration) on ARB and ARGs removal. In addition, the performance of conventional disinfection methods (ozonation, chlorination, UV radiation, and peracetic acid) is investigated. Finally, the application of the widely used Advanced Oxidation Processes (AOPs) for the control of AMR in wastewater is discussed. It is envisaged that this review will provide valuable knowledge for the most efficient technologies for ARB and ARGs removal in WWTP.

2. Bibliometric Analysis

Scopus was used for the retrieval of publications. The treatments studied were selected based on a bibliometric analysis. The search was based on terms related to the elimination of antibiotic resistance from wastewater. An amount of 552 articles were found, limited to English-language articles and reviews from 2009–2022. VosViewer (version 1.6.19) was used to analyze the co-occurrence of author keywords in order to gain insight into the main treatment methods used to remove ARB and ARGs. The co-occurrence network of author keywords is shown in Figure 1. Each cluster represents one author keyword. The larger the cluster, the higher the frequency of a keyword. In addition, the distance between the clusters indicates the relationship between the keywords. The smaller the distance between the clusters, the higher the correlation [37]. As it can be seen, disinfection, advanced oxidation processes, tertiary treatment, membrane bioreactor, and adsorption appear as author keywords. In addition, the main keyword is antibiotic resistance genes (ARGs), which shows that the study of their removal from wastewater is more extensive than antibiotic-resistant bacteria (ARB). A descriptive analysis of annual publications from 2009 to 2022 was also performed using Bibliometrix (version 4.1). The annual growth of publications is 39.32%, and as can be seen, the number of publications published increased significantly in 2018. The increasing publications show that the topic of antibiotic resistance is increasingly becoming the focus of the scientific community.

3. Tertiary Wastewater Treatment Methods

Tertiary wastewater treatment is an important step in obtaining high quality water with a low content of bacteria, pathogens, and suspended solids following secondary treatment [38]. Tertiary treated and disinfected wastewater can be used for crop irrigation and industrial purposes. Tertiary treatment methods include membrane filtration, sand filtration, and granular activated carbon filtration. This review is presenting the current knowledge on the efficiency of these tertiary processes for ARGs and ARB removal from wastewater. It should be underlined however, that there is a lack of scientific literature on the effectiveness of sand filtration and granular activated carbon in removing antibiotic resistance determinants.

3.1. Membrane Filtration

Membrane filtration is widely applied for wastewater reuse [39,40]. According to the pore size, a distinction is made between microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO). The pore size of MF membranes ranges from 0.1 µm to 0.4 μm, the pore size of ultrafiltration is between 0.01 and 0.02 μm, and the pore size of nanofiltration is between 0.5 and 2 nm [41]. Reverse osmosis filters have a pore size between 0.2 and 1 nm.

In Membrane Bioreactors (MBR), microfiltration and ultrafiltration membranes are used for the retention of bacteria, micropollutants, and suspended solids [42]. Major retention mechanisms are size exclusion, electrostatic repulsion, and adsorption of pollutants on membrane surface [43]. Katarzyna Slipko et al. [43] observed that negatively charged membranes had a lower retention of free DNA molecules compared to neutral membranes, which they attributed to the negative charge of DNA. On the contrary, Cristóvão et al. [44] found that negatively charged membranes at a neutral pH retained efficiently blaKPC, bla_OXA-48, bla_NDM, bla_IMP and bla_VIM, qnrA, qnrB, qnrS from wastewater.

Hiller et al. [45] observed that ARGs removal with membrane filtration varies among different ARGs. For instance, vanA showed low removal by UF and MF due to the fact that it is a common gene in mobile genetic elements, such as plasmids that are too small to be retained by membrane pores. Ultrafiltration is able to retain most of ARB presented in wastewater since they have a larger size than the pore size of a membrane [46]. Michael et al. (2022) [46] observed that UF treating activated sludge effluent achieved a 3–4 log removal of enteric opportunist pathogens and significant ARB removal. However, the presence of ARGs in the permeate was between 3.84 × 10¹ to 2.13 × 10⁴ genes copies/100 mL. Nanofiltration proved to be an effective barrier to ARGs in treated wastewater, removing 99.99% of blaKPC, bla_OXA-48, bla_NDM, bla_IMP, bla_VIM, qnrA, qnrB, and qnrS from wastewater effluent samples [47].

A major problem with membrane filtration is the treatment of the concentrate, which may contain a significant amount of ARGs and ARB [40,48]. Hembach et al. [49] observed that the abundance of ARGs decreased by 6–7 log by UF, while their abundance in the concentrate increased by 2 log. They suggested that recirculation of the concentrate to the activated sludge tank may increase HGT between bacterial communities and therefore an inactivation step is required. Krzeminski et al. [40] proposed that UV irradiation can effectively inactivate ARGs in a concentrate downstream of membrane filtration for water treatment.

MBR (Membrane Bioreactor) is a combination of membrane filtration and biological suspension reactor for the removal of organic matter, micropollutants and bacteria [50]. Le et al. [51] found that MBR completely removed ARB and significantly reduced ARGs up to 7.1 log. However, some ARGs were presented in the MF permeate, suggesting that they are part of viruses, bacteriophages, bacteria smaller than the pore size of the membrane or they are in free DNA. In addition, Karaolia et al. [52] observed that MBR did not effectively remove ARGs, and they applied solar-photo Fenton downstream. Furthermore, while MBR efficiently decreased some ARGs by 4.61 log, they were still present in the effluent [53]. In addition, the MBR operating conditions could be critical for the removal of ARGs from wastewater. For instance, Li et al. [54] found that the plant with the longest SRT (27 days) had 2–4 times higher removal of sulI, sulII, and intI compared to other plants (20.5, 16.4, 16.6 days) when studying ARGs removal in full-scale MBRs. They attributed the higher SRT to lower biomass production and, thus, an abundance of ARGs, considering vertical gene transfer as an important factor in ARGs spread. A summary of the available literature on the removal of ARB and ARGs by membrane filtration is shown in

Table 1. Removal of ARB and ARGs by membrane filtration.

Membrane	Matrix	ARB/ARGs	Removal	Reference
MF (450 nm, 22 m2)	Secondary wastewater effluent	sulI ermB vanA	2.1 log <LOD 1 1.1 log	[45]
UF (20 nm, 80 m2) Pilot scale	Secondary wastewater effluent	sulI ermB vanA	2.9 log <LOD 1.2 log	[45]
UF (0.03 μm, 0.2 m2) 1 m3/day	Conventional activated sludge effluent	blaTEM sulI blaOXA	2.4 log 2.4 log 3 log	[46]
UF (10kDA) and NF (150 DA) membranes Lab scale	Biofilm effluent Activated sludge effluent	Tetracycline, Ampicillin, Trimethoprim/sulfamethoxazole, Ciprofloxacin-resistant E. coli	<10 CFU/mL	[55]
Desal 5 DK nanofiltration membrane (54 cm2) 230 L/(m2·h)	Discharged effluent after biological treatment	blaKPC, blaOXA-48, blaNDM, blaIMP, blaVIM, qnrA, qnrB, qnrS	99.99%	[47]
5 DK membranes (7.9 m2) Pilot scale	After biological treatment	blaKPC, blaOXA-48, blaNDM, blaIMP and blaVIM, qnrA, qnrB, qnrS	99.99%	[44]
UF (80 m2, 20 nm) 70 L/m2h UF (25 nm) 35 L/m2h	After biological treatment	blaTEM, tetM, sulI, blaCTX-M, blaCTX-M-32, blaOXA-48	6–7 log	[49]
Fe(VI)-UF Fe(VI): 9 mg/L UF: 150 kDA MWCO	Secondary wastewater effluent	tetM, tetO, tetW, sulI, sulII, ermB, ermF, aac(6′)-Ib-cr, qnrS	3.26–5.01 log	[56]
MBR HRT: 6.1–7.7 h SRT: 14–28 d Microfiltration	Modified Ludzack-Ettinger (MLE) tanks	amikacin, meropenem, ceftazidime, clindamycin, erythromycin, ciprofloxacin, co-trimoxazole, trimethoprim:sulfamethoxazole, tetracycline, vancomycin and chloramphenicol- resistant bacteria	5 to 7.1 log	[51]
MBR HRT:6.1–7.7 h SRT: 14–28 d Microfiltration	Modified Ludzack-Ettinger (MLE)	aac(6′)-Ib, blaKPC, blaCTX-M, blaSHV, blaNDM1, cfr, ermB, mefAE, sulI, sulII, dfrA, tetM, teO, vanA, intI, qnrA, qnrB	1.3–7.1 log	[51]
MBR HRT: 13.1 h SRT: 27 d Membrane type: Hollow fiber, 18.89 L/(m2·h)	MBR dowstreastream of anaerobic-anoxic-oxic (AAO) process	blaTEM, ermB, tetW, tetO, sulI, sulII, addD, and qnrS	3.2–7.3 log	[54]
MBR HRT: 9.8 h SRT: 21 days A hydrophilic polyvinylidene fluoride (PVDF) hollow fiber membrane (0.1–0.4 μm, 5520 m2)	A2 /O-MBR	tetG, tetW, tetX, sulI	0.67–4.73 log	[53]

Note: ¹ LOD: limits of detection.

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3.2. Sand Filtration

Sand filters are used as tertiary treatment of wastewater. Moreover, sand filters can be combined with conventional disinfections methods in order to reduce the dose of disinfectants and the formation of toxic disinfection by-products [57,58]. There are many studies that observed the efficient removal of pathogens by sand filters. Qian et al. [59] showed that E. coli was removed by biological sand filter by 96.1%. Increasing the sand depth, bacteria removal increases due to the increase of adsorption active sites. Langenbach et al. [60] found that slow sand filters removed E. coli and Enterococcus by 98.6–99.8% and 98.9–99.9%, respectively. On the surface of slow sand filters, a biofilm is formed, which has a biological activity and can therefore remove bacteria, organic matter, particles, and nutrients [61,62,63]. Sun et al. [62] observed that slow filtration with aerobic heterotrophic biofilm exhibited higher removal of ARGs than slow filtration without any biofilm. Particularly, the removal of tetA, tetW, sulI, sulII by sand filtration with aerobic heterotrophic biofilm was 2.50, 2.96, 1.92, 2.3 log, respectively. They attributed the enhanced removal of ARGs to the increase of Proteobacteria and Actinobacteria, which are host of sul and tet genes, accordingly. Similarly, Luprano et al. [64] found that the removal of tetA, ermB, sulI, and sulII genes by sand filtration was increased by 0.9–1.1 log after advanced biological treatment. Lüddeke et al. [65] examined the removal efficiency of antibiotic-resistant bacteria by ozonation in combination with slow sand filtration in a pilot scale. Sand filtration downstream of a ozonation reactor causes a further reduction of antibiotic resistant E. coli, enterococci, and staphylococci by a 0.8–1.1 log. However, McConnell et al. [23] found only a slight decrease (0.26 log) in total ARGs (qnrS, sulI, tetO, ermB, sulII, bla_CTX-M, bla_TEM, mecA) with sand filtration following an aerated lagoon. Moreover, Sabri et al. [15] noted that sand filtration following an activated sludge reactor, and a sludge settling tank decreased further ARGs by 0.71–1.75 log.

Hayward et al. [66] observed that lateral flow sand filtration downstream of a septic tank was able to remove ARGs by 2.9 to 5.4 log. In addition, the concentration of qnrS, bla_TEM, and mecA genes reached levels below the limits of detection. In conclusion, there are very few studies investigating the application of sand filtration for the inactivation of ARB and removal of ARGs in secondary treated wastewater.

3.3. Granular Activated Carbon (GAC)

Activated carbon is often used for the adsorption of various pollutants from wastewater such as organic compounds, heavy metals, and pharmaceutical products [67,68]. Activated carbon can be used for the adsorption of antibiotics, ARGs, and ARB found in wastewater [69]. However, ARB and ARGs removal has not yet been widely studied. Activated carbon is often used as post treatment of ozonation for the removal of possible ozone by-products. There are not many studies that clarify if activated carbon increases or decreases ARB and ARGs. Slipko et al. [70] used a granular activated carbon filter downstream of a pilot-scale ozonation system and observed that sulI and tetW genes were reduced by up to 1 log while the relative abundance of bla_TEM increased significantly. Spit et al. [69] observed in a pilot-scale ozonation and GAC filtration system that vancomycin-resistant enterococci (VRE) increased by 0.5 log after GAC filtration, demonstrating that activated carbon increases antibiotic resistance. Moreover, they contributed the augment of VRE to acquiring of ARGs released by cell lysis of bacteria caused by ozonation. Yang et al. [16] observed that the removal efficiency of GAC filter column operated in continuous mode for 96 h was higher than ozonation (60 mg/L of ozone and 20 min of contact time). Specifically, GAC filter reduced tetO, tetW, and bla_CTX-M genes by 0.53–1.73 log. When ozonation combined with amorphous GAC filter was employed, the tetO, tetW, and bla_CTX-M genes were completely eliminated, and sulI and sulII genes were removed by 2.63–2.75 log. They concluded that GAC filters can be used following ozonation for the removal of extracellular ARGs.

3.4. Special Additivities for ARB and ARG Control

Nanomaterials have the potential of inactivating microorganisms found in wastewater [71]. Nanomaterials can be a promising disinfectant tool due to their strong redox, absorption, and photocatalytic activity [72]. Nanoparticles (NPs) are a more economical alternative treatment method in comparison with AOPs [73]. NPs have a size ranging from 1 to 100 nanometers (nm) and possess unique properties [74]. Due to their unique electronic structure, NPs have the ability to bind DNA through electrostatic interactions and facilitate the insertion of DNA onto the surface of oxide nanoparticles using chaotropic salt [73]. As a result, they could serve as a useful method for removing ARGs. NPs have antibacterial activity through several pathways, including interactions with DNA and proteins, oxidative stress through the formation of ROS, interactions with cell wall components, and prevention of biofilm formation [75]. All these properties make nanoparticles a promising tool to combat multidrug-resistant bacteria. Yu et al. [76] found that sulfided nano-zerovalent iron combined with persulfate (S-nZVI/PS) was effective for the removal of ARB and ARGs from wastewater effluents. Specifically, S-nZVI/PS could completely inactivate ARB in 10 min and extracellular ARGs in 5 min. However, S-nZVI/PS reduced intracellular bla_TEM and tetA in 60 min. In addition, the graphene oxide nanosheet (GO) was able to remove ARGs below the detection limit at a concentration of 500 μg/mL mainly due to the presence of oxygen-containing groups and a π-binding system on the GO nanosheet. These characteristics enable chemical binding with aromatic nucleic acids and facilitate strong π-stacking interaction [76]. Another promising material for the removal of ARB and ARGs may be copper oxide nanoparticles. Eid et al. [77] found that plant-based copper oxide nanoparticles have antimicrobial activity against pathogens (Staphylococcus aureus, Bacillus subtilis, Escherichia Coli, Pseudomonas aeruginosa, and Candida albicans). The mechanism of inhibition is attributed to the decay of CuO-NPs into toxic Cu²⁺ ions and the overproduction of ROS. Hamad et al. [78] proposed that nZVI and copper oxide nanoparticles (CuO-NPs) are environmentally friendly adsorbents for ARGs. They found that nZVI reduced β-lactamase genes from 47.2% to 94% and fluoroquinolone genes from 34% to 89.4% in hospital wastewater. CuO-NPs reduced β-lactamase genes from 34.4% to 89.4% and fluoroquinolone genes from 43.3% to 77.1%. In addition, Rajapaksha et al. [79] found that copper oxide nanoparticles (CuO-NPs) doped in graphene oxide (GO) have great potential to inhibit the growth of E. coli and S. typhimurium bacteria. The CuO-based nanomaterials can produce ROS, which can degrade the cell wall, resulting in cell lysis and death. Ezeuko et al. [80] observed that silver nanoparticles (AgNPs) were effective in removing bacterial DNA conveying ARGs from water. The antibacterial properties of nanomaterials could be a useful tool to prevent the spread of ARB and ARGs in wastewater, but further studies are needed.

4. Wastewater Disinfection Methods

Wastewater disinfection is a critical step for the protection of environment and public health and the safe reuse of wastewater. The most common wastewater disinfection methods are chlorination, ultraviolet (UV), and ozonation [81]. An alternative disinfection method to chlorination is peracetic acid (PAA) due to the low formation of disinfection by-products (DBPs) [82]. However, some researchers showed that plasmids remain functional after PAA disinfection while chlorination has more permanent damage to plasmids [83]. The permanent damage of plasmids is crucial since they are often carriers of ARGs, facilitating HGT. In this section, the capability of chlorination, UV, ozonation, and PAA for the removal of ARB and ARGs is discussed.

4.1. Chlorination

Chlorination is the most common disinfection technique in the world that prevents the spread of numerous waterborne diseases in treated water and wastewater. Chlorination is known to irreversibly damage the cellular surface of bacteria, the membrane permeability, the nucleic acid, and bacterial enzymes; therefore, it causes the inactivation of bacteria [84,85,86]. However, Zheng et al. [85] reported that chlorine may not effectively degrade bacterial DNA, and therefore, the removal of ARGs depends to a great extent on the inactivation of bacteria. Numerous studies have shown that higher doses of chlorine are needed for the removal of ARGs compared with the inactivation of ARB [87,88]. Liu et al. (2018) [89] found that chlorination increased the concentration of intracellular and extracellular ARGs in secondary wastewater effluent, showing that an insufficient chlorine dose can spread antibiotic resistance in non-antibiotic resistant bacteria in the environment. ARGs can survive without the presence of their hosts, and thus, there is a potential risk of their transfer to a new host, enhancing the horizontal gene transfer [90]. Wang et al. [91] observed that the relative abundance of ARGs in WWTP-treated hospital wastewater increased from 0.22 to 2.23 log after disinfection with sodium hypochlorite. Moreover, during chlorination, naked DNA, ARGs, and mobile genetic elements are released from the dead ARB in the wastewater, posing the risk of spreading the antibiotic resistance among organisms in the environment. Jin et al. [92] found that chlorination enhanced the horizontal transfer of plasmids to other bacterial communities by natural transformation and that higher doses of chlorine were necessary in order to eliminate them. Hou et al. [93] observed that the antibiotic resistance to ceftazidime, chloramphenicol and ampicillin increased by 1.4–5.6 fold in cultivable chlorine-injured P. aeruginosa due to the oxidative stress. Moreover, many studies found the enrichment of ARB due to the chlorination [94,95]. Chlorination has different elimination potential for different antibiotic-resistant bacteria. Yuan et al. [90] found that sulfadiazine- and erythromycin-resistant heterotrophic bacteria were tolerant to low doses of chlorine and required a higher initial dose to become inactivated. Moreover, they found that low doses of chlorine (<40 mg min/L) increased the conjugative transfer of ARGs. During low doses of chlorination, chlorine reacted with ammonia nitrogen, and the formation of chloramine occurred. The formation of chloramine enhanced the production of the pilus on the surface of conjugative cells that amplified the plasmid transfer between the cells. At high chlorine doses (>80 mg min/L), conjugative transfer was not feasible because the concentration of antibiotic-resistant E. coli was low (<10⁴ CFU/mL), and thus, there were less opportunities for bacteria to come in contact. A summary of the available literature on the inactivation of ARB and ARGs by chlorination is presented in

Table 2. Inactivation of ARB and ARGs by chlorination.

Matrix	Dose	ARB/ARGs	Removal Efficiency	Reference
Conventional wastewater treatment	8 mg Cl2/L, t = 30 min	tetA, tetM, tetO, tetQ and tetW	>85%	[85]
	16 mg Cl2/L, t = 30 min	sul genes	>99%
	32 mg Cl2/L, t = 30 min	tetracycline, sulfamethoxazole and multiple antibiotic-resistant bacteria	100%
MBR effluent	2 mg/L total chlorine, t = 4 min	multidrug-resistant E. coli	5 log	[96]
Conventional wastewater treatment	0.5 mg/L total chlorine, t = 30 min	vancomycin-, cephalexin-, rifampicin-, gentamicin-, tetracycline-, ciprofloxacin-, chloramphenicol-resistant bacteriamultidrug-resistant bacteria	>9 log	[90]
		ere(A), erm(B)	87%, 40%
Conventional wastewater treatment	10 mg/L total chlorine, t = 180 min	sulI, tetM, qnrS, blaTEM	8% for sulI, tetM, qnrS and 4% for blaTEM	[97]
Activated sludge treatment	30 mg/L Cl2, t = 30 min	tetX, tetG, sulI	1.2–1.5 log	[98]
Activated sludge + chemical treatment	0.05 mg/L residual chlorine, t = 45 min	tetA, ermB, blaTEM	>99%	[22]
Activated sludge treatment	1 mg/L total chlorine, 0.2 mg/L residual chlorine, t = 15 min	multidrug resistant E. coli	100%	[99]

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4.2. Ozonation

Ozone is a strong disinfectant with an oxidation potential of 2.7 V. Hydroxyl radicals and reactive oxygen species are formed during the decomposition of ozone under alkaline conditions [100,101,102]. Ozone attacks the bacterial cell wall and causes cell death. Molecular ozone damages the cell membrane, causing the oxidation of cellular components, breaking single strands, damaging bases, and altering plasmid conformations [102,103,104]. Ozone doses higher than 0.5 gO₃/gDOC have been found to be effective for bacterial inactivation [70,105]. K. Slipko et al. [70] suggested that an ozone dose higher than 0.6 gO₃/gDOC is required to inactivate ARGs and ARB.

Iakovides et al. [106] observed complete inactivation of fecal coliforms, Enterococcus, E. coli, and Pseudomonas aeruginosa resistant to ofloxacin, trimethoprim, erythromycin, and sulfamethoxazole, respectively, in laboratory-scale experiments using ozonation over conventional activated sludge effluent and an ozone dose of 1.5 gO₃/gDOC and a contact time of about 10 min. Inactivation of most of these ARB from MBR wastewater occurred at a lower ozone dose of 1 gO₃/gDOC due to their lower initial concentrations. A longer contact time was required for complete inactivation of ARGs in secondary wastewater effluent. Specifically, inactivation of the dfrA1 gene occurred in 120 min and that of sulI and aadA1 in 180 min at an ozone dose of 1.5 gO₃/gDOC. In MBR-treated wastewater, the ARGs studied were inactivated in only 30 min and at a lower ozone dose of 1 gO₃/gDOC. However, the inactivation with ozone was not permanent, as the abundance of ARGs increased after 72 h of storage in the dark and at room temperature. The proliferation of the genes could be due to the regrowth of bacteria and to activated sludge flocs acting as a protective barrier for bacteria. The results are consistent with another study by Iakovides et al. [107], where they found that qacDE1, sulI, aadA1, and dfrA1 genes increased after 72 h of continuous treatment and reached levels similar to the untreated secondary effluent. Consistent with these results, Sousa et al. [108] observed that although ozonation removed ARGs below the limit of quantification, bla_TEM and sulI genes reached levels close to pre-treatment levels after three days of storage. It appears that ozonation causes temporary rather than permanent damage, and bacteria can be reactivated when oxidative stress is relieved. Yang et al. [16] reported that increasing the ozone dose from 10 to 60 mg/L did not contribute to a higher removal efficiency rate. A higher ozone dose was able to inactivate bacteria protected by particles and flocs, causing intracellular ARGs to enter the effluent and to increase the abundance of ARGs. However, the higher ozone dose of 60 mg/L was effective in efficiently removing ARGs from secondary treated wastewater.

The disinfection ability of ozone varies greatly among different bacterial species and depends on numerous factors, such as the cell envelope, growth, and repair [109]. For instance, Alexander et al. [110] observed that enterococci exhibited higher robustness towards ozone compared to Pseudomonas aeruginosa. In addition, K. Slipko et al. [70] observed that the inactivation of ampicillin-resistant E. coli required a higher ozone dose than trimethoprim- and sulfamethoxazole-resistant E. coli.

At full scale, ozone does not effectively remove ARGs from secondary effluents, possibly due to consumption of ozone by dissolved organics [109,111]. In addition, Alexander et al. [110] observed the increase of vanA and bla_VIM in abundance with surviving bacteria by 4- and 7-fold after ozonation treatment in a pilot scale ozone system. Increasing the ozone dose enhances the ARB removal efficiency [69,70,107]. However, Zheng et al. [85] observed that increasing the ozone dose did not contribute to a significantly higher ARGs removal efficiency. They found that ozone causes bacterial lysis that results in the release of ARGs in the wastewater solution, and as the concentration increases, more ozone is required to effectively remove ARGs. A summary of ARB and ARGs inactivation with ozone disinfection is shown in

Table 3. ARB and ARGs inactivation with ozone.

Matrix	Conditions	ARB/ARGs	Removal Efficiency	Reference
Conventional activated treated sludge	1.5 gO3/gDOC, t = 10 min	fecal coliforms, Enterococcus, E. coli, and Pseudomonas aeruginosa resistant to ofloxacin, trimethoprim, erythromycin, and sulfamethoxazole	<LOD	[106]
	1.5 gO3/gDOC, t = 120–180 min	sulI, aadA1, dfrA1	<LOD
MBR treated effluent	1 gO3/gDOC, t = 2–6 min	fecal coliforms, Enterococcus, E. coli, and Pseudomonas aeruginosa resistant to ofloxacin, trimethoprim, erythromycin, and sulfamethoxazole	<LOD	[106]
	1.5 gO3/gDOC, t = 120–180 min	sulI, aadA1, dfrA1	<LOD
Conventional activated treated sludge	0.3 gO3/gDOC t = 15 min	erythromycin-resistant E. coli	Total inactivation	[112]
Secondary treated effluent following a pilot scale ozone system	0.9 gO3/gDOC, HRT: 18 ± 2 min	ermB	99.3%	[110]
Secondary treated effluent	0.38 gO3/gDOC, t = 25 min	Extended-Spectrum Beta-Lactamase (ESBL) producing E. coli and Vancomycin Resistant Enterococci (VRE)	1 log	[69]
Secondary treated effluent	2 mg/L O3, t = 10 min	tetracycline and sulfamethoxazole-resistant bacteria	2.10–2.46 log	[85]
Secondary clarifier effluent	0.55 g O3 g/gDOC	bacteria and bacteria multiresistant to sulfamethoxazole/trimethoprim/tetracycline and norfloxacin/ceftazidime	1.4−1.6 log	[109]
Secondary treated effluentBench-scale continuous mode ozonation	0.75 g O3 g/gDOC HRT: 40 min	qacDE1, sulI, aadA1, dfrA1	4–6 log	[107]
	0.25 gO3/gDOC HRT: 40 min	sulfamethoxazole and trimethoprim- resistant E. coli	Total inactivation
Conventional Activated sludge and ozone reactor	0.40 gO3/gDOC HRT: 9–40 min	trimethoprim and sulfamethoxazole resistant E. coli	1.99 log	[70]
	0.90 gO3/gDOC HRT: 9–40 min	ampicillin-resistant E. coli	2.50 log
Secondary treated effluent	10 mg/L O3, t = 20 min	tetA, tetO, tetW, sulI, sulII	0.16–0.50 log	[16]
	60 mg/L O3, t = 20 min	tetA, tetO, tetW, sulI, sulII	80–90%
Secondary treated effluent	50gN/m3, t = 30 min	blaTEM, qnrS, sulII, vanA	<LOD	[108]
Effluent from biological treatment-Full scale	1 gO3/gDOC Flow rate: 7 m3/h	blaTEM, tetM, sulI, CTX-M, CTX-M-32, blaOXA-48	2 log	[49]

.

4.3. Ultraviolet Irradiation

UV irradiation in the UV-C band (250–270 nm) is one the most applied disinfections methods in wastewater treatment due to the non-direct formation of disinfection by-products and the short contact time of application [8,113].

UV irradiation has different elimination potential for different antibiotic-resistant bacteria and genes [88] (Table 4). Many studies have shown that the abatement of ARGs increases with the increase of UV fluence [85,114]. Increasing the UV dose from 80 to 400 mJ/cm² increased the abundance of tetB, sulII, and tetA by 1.8-, 36.3-, and 53.4-fold, respectively [115]. UV disinfection (30 and 100 mWs/cm²) did not reduce the concentrations of tetQ and tetG genes in treated effluent in a full-scale wastewater treatment plant [116]. This is consistent with the study of Rafraf et al. [117], in which a UV system was unable to remove the bla_TEM, qnrA, and sulI genes in activated sludge secondary effluent. McKinney and Pruden [118] found that inactivation of ARGs requires an order of magnitude higher UV fluence (200–400 mJ/cm² for a 3- to 4-log reduction) than inactivation of ARB (10–20 mJ/cm² for a 4- to 5-log reduction). Yingying Zhang et al. [98] found that UV alone is unable to remove ARGs from wastewater, with a log reduction of ~0.1–0.23 for a UV dose of 62.4 mJ/cm². Ping et al. [119] also found an increase in tetracycline, sulfonamide, quinolone, macrolide, and β-lactam resistance genes from 0.6 to 30.6% at a UV dose of 36 mJ/cm² due to the promotion of the conjugative transfer of ARGs at low UV doses. The presence of ARB and antibiotics after UV disinfection might cause the spread of ARGs [4,64]. In addition, L. Yang et al. [16] reported that a UV unit (20 mJ/cm²) at a full scale WWTP was not able to remove tetA, tetO, tetW, sulI, sulII, bla_CTX-M genes.

In addition, UV irradiation could increase antibiotic resistance. Zheng et al. [85] observed that the percentage of tetracycline- and sulfamethoxazole-resistant bacteria increased at low UV fluence, indicating tolerance to UV. Specifically, the proportion of the former increased from 3.3% to 7.08% while the proportion of the latter increased from 29.4% to 52.9% at a UV fluence of 20 mJ/cm². A higher UV fluence of 80 mJ/cm² was required for complete inactivation. Similar results were also reported by Guo et al. [114] that found that the percentage of tetracycline-resistant bacteria almost doubled after UV disinfection at a low fluence of 5 mJ/cm². In addition, Huang et al. [120] observed that the percentage of tetracycline-resistant bacteria increased significantly from 1.65% to 15.52% when the UV dose reached 20 mJ/cm². It is worth noting that the removal of the total heterotrophic bacteria was 4 log at 20 mJ/cm² while the removal of tetracycline-resistant bacteria at the same UV dose was 3 log. Meckes [121] observed the increase of tetracycline-resistant bacteria was due to the protein tet, which absorbs irradiation at 254 nm and protects the bacteria from UV radiation. All these results show that the tetracycline-resistant bacteria are tolerant to UV disinfection. Multidrug-resistant E. coli were found to be tolerant to a low dose of UV, and a dose of ≥20 mJ/cm² was required to initiate their elimination, which is due to the fact that multidrug-resistant bacteria carry more ARGs and can survive longer [115]. It seems that ARB need a higher dose of UV for their complete inactivation.

It is known that bacteria can repair their DNA after UV irradiation under light and dark conditions due to photoreactivation and dark repair [122]. The enzyme photolyase binds to CPD during photoreactivation and uses the energy of radiation (320–500 mm) to repair DNA damage. Under dark conditions, damaged DNA is replaced with intact nucleotides via multi-enzymes [123,124]. Photoreactivation poses significant risks when the treated effluents are discharged into surface waters, as there is a possibility that bacteria will be reactivated under solar radiation [125]. Dark repair can take place when the treated wastewater is stored to be reused.

Tetracycline-resistant isolated strains and total heterotrophic bacteria increased substantially after irradiation with 20 mJ/cm² or 40 mJ/cm² followed by incubation for 22 h under dark conditions at 25 °C [120]. In addition, Childress et al. [126] observed that tetracycline-resistant heterotrophs were reactivated and reached higher concentrations after 48 h of UV treatment under dark and light conditions. Sousa et al. [108] found that the concentrations of bla_TEM and qnrS after three days of storage reached almost the same values as before disinfection.

Table 4. Inactivation of ARB and ARGs by UV irradiation.

Treatment Prior to UV	Dose	ARB/ARGs	Removal Efficiency	Reference
Conventional wastewater treatment	40 mJ/cm2	tetA, tetM, tetO, tetQ and tetW,	52.0–73.5% 79.7–92%	[85]
	160 mJ/cm2	sulI, sulII	78.1%, 71.1%
Activated sludge treatment	249.5 mJ/cm2, Irradiation time: 60 s	tetX, sulI, tetG	0.58, 0.30–0.40 log	[98]
Activated sludge +chemical treatment	80 UV-C lamps 25.8 mJ/cm2, Irradiation time: 2 s	tetO, ermB, qnrS	93.7–99%	[22]
Secondary Treated effluent	240 mJ/cm2	macrolides-resistant bacteria sulfonamides-resistant bacteriatetracyclines-resistant bacteriaquinolones-resistant bacteria	86.5% 89.5% 89.9% 99%	[127]
After biological aerated filter process in	5 mJ/cm2	heterotrophic bacteria resistant to erythromycin and tetracycline	1.4 ± 0.1 and 1.1 ± 0.1, respectively	[114]
		ereA, ereB, ermA and ermB	3.0 ± 0.1 log
		tetA, tetB, tetM and tetO	1.9 ± 0.1 log
Secondary treated effluent	5 mJ/cm2	tetracycline resistant bacteria	3 log	[120]
Secondary treated effluent	5 mJ/cm2	tetB, sulII	1.18, 0.85 log	[115]
Secondary treated effluent	10–20 mJ/cm2	methicillin-resistant Staphylococcus, vancomycin-resistant Enterococcus faecium, Escherichia Coli SMS-3–5	4–5 log	[118]
Secondary treated effluent	200–400 mJ/cm2	mecA, vanA, tet(A), ampC	3–4 log	[118]
Cyclic activated sludge effluent	20 mJ/cm2 Full scale	tetA, tetO, tetW, sulI, sulII, blaCTX-M,	0.27–0.40 log	[16]

Table 4. Inactivation of ARB and ARGs by UV irradiation.

Treatment Prior to UV	Dose	ARB/ARGs	Removal Efficiency	Reference
Conventional wastewater treatment	40 mJ/cm2	tetA, tetM, tetO, tetQ and tetW,	52.0–73.5% 79.7–92%	[85]
	160 mJ/cm2	sulI, sulII	78.1%, 71.1%
Activated sludge treatment	249.5 mJ/cm2, Irradiation time: 60 s	tetX, sulI, tetG	0.58, 0.30–0.40 log	[98]
Activated sludge +chemical treatment	80 UV-C lamps 25.8 mJ/cm2, Irradiation time: 2 s	tetO, ermB, qnrS	93.7–99%	[22]
Secondary Treated effluent	240 mJ/cm2	macrolides-resistant bacteria sulfonamides-resistant bacteriatetracyclines-resistant bacteriaquinolones-resistant bacteria	86.5% 89.5% 89.9% 99%	[127]
After biological aerated filter process in	5 mJ/cm2	heterotrophic bacteria resistant to erythromycin and tetracycline	1.4 ± 0.1 and 1.1 ± 0.1, respectively	[114]
		ereA, ereB, ermA and ermB	3.0 ± 0.1 log
		tetA, tetB, tetM and tetO	1.9 ± 0.1 log
Secondary treated effluent	5 mJ/cm2	tetracycline resistant bacteria	3 log	[120]
Secondary treated effluent	5 mJ/cm2	tetB, sulII	1.18, 0.85 log	[115]
Secondary treated effluent	10–20 mJ/cm2	methicillin-resistant Staphylococcus, vancomycin-resistant Enterococcus faecium, Escherichia Coli SMS-3–5	4–5 log	[118]
Secondary treated effluent	200–400 mJ/cm2	mecA, vanA, tet(A), ampC	3–4 log	[118]
Cyclic activated sludge effluent	20 mJ/cm2 Full scale	tetA, tetO, tetW, sulI, sulII, blaCTX-M,	0.27–0.40 log	[16]

4.4. Peracetic Acid

Peracetic acid (PAA) is an effective wide spectrum disinfectant for inactivating bacteria, viruses, and protozoa in wastewater. PAA contains 5 to 15% w/w peracetic acid, hydrogen peroxide, acetic acid, and water [128]. The main advantage of PAA compared to chlorine is the limited formation of mutagenic and toxic DBPs [129,130]. PAA inactivates bacteria by oxidizing sulfhydryl sulfur bonds in proteins, metabolites, and enzymes of cell wall and by disrupting cell membranes. Through cell wall rupture, PAA interferes with the lipoprotein cytoplasmic membrane’s chemiosmotic function and transport [131,132]. Moreover, bacteria are also inactivated by the release of hydroxyl radicals [133].

A summary of PAA efficiency on ARB and ARGs inactivation is shown in Table 5. Chhetri et al. [134] found that PAA was effective in removing ciprofloxacin-resistant bacteria from municipal and hospital wastewater. Removal efficiency increases with an increasing contact time and concentration of PAA. Campo et al. [135] found that peracetic acid was able to control antibiotic resistance in secondary treated wastewater by reducing the percentage of bacteria resistant to ampicillin from 40% to 2.7% at 3 mg/L PAA and a contact time of 16 min. Ping et al. [119] observed the increase in tetracycline, sulfonamide, quinolone, macrolide, and β-lactam resistance genes from 13.4 to 40.6% at a PAA concentration of 8 mg/L and a contact time of 24 min, and they attributed this to the permeability of the bacterial cell membrane. Under oxidative stress, bacteria express outer membrane proteins, which increases membrane permeability and promotes the spread of ARGs via HGT. In addition, they observed that the major host bacteria of tetracycline resistance genes were Bacteroides, Prevotella, and Streptococcus, which increased after PAA disinfection, which was highly related to the increase in tetracycline resistance genes due to mutations, HGT, and stimulation of the multidrug efflux system. Luprano et al. [64] found that disinfection of treated wastewater with 1 mg/L PAA and that a contact time of 30 min was not effective for removing ermB, sulI, and tetA genes while sulII was only slightly reduced by 0.4 log.

Balachandran et al. [96] found that a PAA dose of 4 mg/L and a contact time of 7 min provided a 4 log reduction in multidrug-resistant E. coli. However, compared with chlorine under the same conditions (2 mg/L disinfectant, pH = 7.5), the inactivation was faster than with PAA. At a contact time of 4 min, the removal of multidrug-resistant E. coli and enterococci was 4 log while no removal was obtained with PAA.

The presence of acetic acid as an organic constituent in wastewater poses a potential risk for bacterial regrowth [131]. Jing et al. [136] found that, at a PAA dose of 5 mg/L and a contact time of 10 min, the concentrations of chloramphenicol- and tetracycline-resistant bacteria after 22 h of PAA disinfection were 10- and 40-fold higher, respectively, than the concentrations in the non-disinfecting secondary effluent. Regrowth was inhibited at a PAA dose of 12 mg/L, indicating that higher PAA doses are required for complete inactivation of ARB. In contrast, Balachandran et al. [96] observed no regrowth after 72 h of PAA disinfection at a dose of 3 mg/L and a contact time of 10 min. There is an urgent need for more scientific research in order to elucidate the effect of PAA on ARGs and on the viable but not cultivable state of bacteria. There is a large knowledge gap regarding the removal of ARGs by PAA in wastewater processes.

Table 5. Inactivation of ARB and ARGs by peracetic acid.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Raw municipal wastewater	25 mg/L PAA t = 60 min	ciprofloxacin-resistant bacteria	1.9 log	[134]
	75 mg/L PAA t = 60 min	ciprofloxacin-resistant bacteria	<LOD	[134]
Secondary treated wastewater	55 mg.min/L	ampicillin-resistant E. coli	Total inactivation	[135]
Secondary treated effluent	20 mg/L PAA t = 10 min	ampicillin resistant bacteriatetracycline resistant bacteria	2.3 log 1.1 log	[136]
Secondary treated effluent	12.5 mg/L PAA t = 10 min	penicillin-, cefalexin-resistant bacteria	2.1 log	[136]
MBR effluent inoculated with multidrug resistant bacteria	3 mg/L PAA t = 7 min 4 mg/L PAA	multidrug resistant E. coli	1.2 log 4 log	[96]
	3 mg/L PAA t = 10 min 4 mg/L PAA	multidrug resistant enterococci	1.2 log 3.5 log	[96]
Secondary treated effluent	8 mg/L PAA t = 24 min	tetA, tetB, tetC, tetM, tetQ, tetW, tetX, sulI, sulII, qnrA, qnrD, qnrS, ereA, ermA, ermB, ermF, mefA, blaCTXM, blaOXA-1 and blaTEM	−13.6% to −40.6%	[119]
Sequencing Batch Biofilter Granular Reactor (SBBGR)	1 mg/L PAA t = 24 min	sulII ermB, sulI, tetA	0.4 log no removal	[64]

Table 5. Inactivation of ARB and ARGs by peracetic acid.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Raw municipal wastewater	25 mg/L PAA t = 60 min	ciprofloxacin-resistant bacteria	1.9 log	[134]
	75 mg/L PAA t = 60 min	ciprofloxacin-resistant bacteria	<LOD	[134]
Secondary treated wastewater	55 mg.min/L	ampicillin-resistant E. coli	Total inactivation	[135]
Secondary treated effluent	20 mg/L PAA t = 10 min	ampicillin resistant bacteriatetracycline resistant bacteria	2.3 log 1.1 log	[136]
Secondary treated effluent	12.5 mg/L PAA t = 10 min	penicillin-, cefalexin-resistant bacteria	2.1 log	[136]
MBR effluent inoculated with multidrug resistant bacteria	3 mg/L PAA t = 7 min 4 mg/L PAA	multidrug resistant E. coli	1.2 log 4 log	[96]
	3 mg/L PAA t = 10 min 4 mg/L PAA	multidrug resistant enterococci	1.2 log 3.5 log	[96]
Secondary treated effluent	8 mg/L PAA t = 24 min	tetA, tetB, tetC, tetM, tetQ, tetW, tetX, sulI, sulII, qnrA, qnrD, qnrS, ereA, ermA, ermB, ermF, mefA, blaCTXM, blaOXA-1 and blaTEM	−13.6% to −40.6%	[119]
Sequencing Batch Biofilter Granular Reactor (SBBGR)	1 mg/L PAA t = 24 min	sulII ermB, sulI, tetA	0.4 log no removal	[64]

5. Advanced Oxidation Processes (AOPs) for Wastewater Treatment

Disinfection alone may not be efficient for the complete removal of ARB, ARGs, and pathogenic bacteria, and therefore, there is an urgent need to combine conventional treatment with oxidants. Advanced oxidation methods are efficient for the removal of pathogens from wastewater [137]. Common oxidants used in AOPs are ozone (O₃), UV, and hydrogen peroxide (H₂O₂). The H₂O₂/O₃, O₃/UV, and H₂O₂/UV procedures are very commonly used. Some of these technologies are widely known and well developed while others are rather new, such as photocatalysis with titanium dioxide (TiO₂) and the Fenton reaction. Advanced oxidation processess have gained significant attraction as a disinfectant tool due to the low production of toxic disinfection by-products compared to the conventional disinfection methods, such as chlorination and ozonation [138,139]. Advanced oxidation processes produce highly reactive oxygen species (ROS), such as a hydroxyl radical (HO·), sulfate radical (SO4·⁻), and superoxide radicals (O₂·⁻) and due to their oxidative potential, can damage the cell membranes and walls, enzymes, proteins, and DNA of pathogens [137,140]. ROS break down the double stands of DNA, and therefore, the removal of ARGs can occurr [141]. Moreover, H₂O₂ has a similar polarity with water and, therefore, can diffuse into the bacterial cells, increasing their permeability and causing high mutation rates [142,143].

5.1. UV-Induced AOPs or Photochemical Advanced Oxidation Processes

UV has been found to have limited potential to reduce ARGs from wastewater [85,114,116], and thus, there is an urgent need to add oxidants to increase removal efficiency. UV-induced AOPs are widely used due to low energy consumption and low formation of by-products, which make them environmentally friendly [141]. Moreover, Wang et al. [91] observed that, under scanning electron microscopy, there was little damage to the bacteria’s surface at a UV dose of 32 mJ/cm² and that high UV doses are required for the permanent destruction of bacteria. The presence of oxidative radicals increases the removal potential of ARB and ARGs from wastewater compared to UV alone [144]. This review covers the efficiency of UV/H₂O₂, UV/chlorine, UV/persulfate induced AOPs to remove ARB and ARGs from wastewater. However, there is an urgent need for more research for the potential of UV-induced AOPs to inactivate ARB and ARGs from wastewater. Moreover, there is also a need for investigation, whereas ARB exhibits photoreactivation and dark repair after the treatment.

5.1.1. UV/H₂O₂

Beretsou et al. [145] observed total inactivation of fecal coliforms, total heterotrophs, Enterococcus, Pseudomonas conferring resistance to trimethoprim, erythromycin, and ofloxacin under the combination of UV dose of 0.1–1 J/cm² and 40 mg/L of H₂O₂. The high dose of H₂O₂ prevented the ARB from self-repair after 24 h, showing permanent oxidative damage due to the presence of hydroxyl radicals. On the contrary, a higher UV dose of 16 J/cm² was required to completely remove ARGs from wastewater, showing their persistence even after the full inactivation of their hosts. Inactivation of ARB leads to the release of ARGs in wastewater, which require increased UV dose for efficient removal. Rodríguez-Chueca et al. [144] found that UV/H₂O₂ did not contribute to the higher ARGs removal in a full-scale tertiary wastewater treatment plant. The removal rate of ARGs by UV-C (42 J/L; 4 s) was 0.41 log, whereas the log removal by UV/H₂O₂ (42 J/L, 4 s, H₂O₂; 0.5 mM) was 0.21 log. They attributed this to a competition for the adsorption of UV photons between oxidants and DNA, resulting in a decrease in photolysis of DNA, thus eliminating the potential for ARGs removal. Ferro et al. [146] found that UV/H₂O₂ effectively removed multidrug-resistant E. coli from urban wastewater after an exposure time of 5 min. Moreover, UV/H₂O₂ efficiently removed bla_TEM and qnrS genes from intracellular DNA. On the other hand, they observed that the percentage of bla_TEM in total DNA increased after 240 min of treatment. During the treatment, dead bacterial cells are released in the solution, which elevates the DNA concentration. They concluded that UV/H₂O₂ treatment may not be effective in preventing the spread of antibiotic resistance in the environment due to the release of free DNA in the suspended solution. Similar findings were reported by Michael et al. [147], after UV/H₂O₂ treatment, some ARGs were persistent, while others were reduced below the limit of quantification. The removal efficiency varies among different ARGs. For instance, Das et al. [148] observed that genes with larger amplicon sizes displayed higher elimination rates due to higher activated sizes for ·OH radicals to attack. Moreover, the UV/H₂O₂ was not able to remove the suII gene due to its smaller amplicon size [145].

5.1.2. UV/Chlorine

The synergistic effect of chlorine and UV can enhance the efficiency of disinfection and reduce the dose of chlorine, thus eliminating the formation of DPBs [91]. There is recent evidence that DBPs could promote the antibiotic resistance via the genetic mutations and horizontal gene transfer of ARGs [149]. Shekhawat et al. [150] showed that the combined UV and chlorine (2.5 mg/ L of chlorine and 27 mJ/cm² UV) significantly decreased the concentrations of total tri-halomethanes (THMs) and total haloacetic acids (THAAs) in wastewater, curbing the potential harmful impacts of their presence in aquatic environments. Chlorination and UV can effectively remove ARB from wastewater effluents, but ARGs cannot be removed on large a scale [151,152]. The combination of UV and chlorine produces hydroxyl radicals (HO·) and radicals of reactive chlorine species (RCS) (Cl·; E_h = 2.4 V, Cl₂·⁻; E_h = 2.0 V, ClO·; E_h = 1.5–1.8 V), and these RCS were found to be responsible for the removal of ARGs [153]. Chlorination and UV may not completely destroy bacteria, but it can cause them to enter the viable but nonculturable state (VNBC), the synergestic effect of UV and chlorine has been found to permanently remove the VBNC cells, eliminating the potential transfer of ARGs [154,155,156]. Yingying Zhang et al. [98] reported that UV/chlorine was more effective in removing ARGs from secondary treated wastewater than UV and chlorine alone. The log reduction by UV/chlorine (UV dose of 249.5 mJ/cm² and chlorine dose of 30 mg/L) for the tetX gene was 2.1 log while the log reduction by individual UV and chlorination was 0.6 and 1.5, respectively. It is noteworthy that with increasing UV and chlorine dose, the removal of ARGs was higher. In addition, Ye et al. [157] found that the relative abundance of ARGs increased after a short treatment with UV/chlorine (1 and 10 min), whereas the relative abundance decreased significantly when the treatment was extended to 30 min. Wang et al. [91] found that the combination of UV (≥4 mJ/cm²) and chlorine (≥1 mg/L) was effective to inhibit the conjugative transfer of plasmid RP4 compared to the standalone UV and chlorination, thus eliminating the dissemination of antibiotic resistance by HGT. Moreover, they observed that the synergistic effect of UV (32 mJ/cm²) and chlorine (1 mg/L) could impair the photoreactivation of ARB to a certain degree. However, bacteria resistant to different antibiotics exhibit different sensitivity to UV/chlorination [88]. There is an urgent need for further literature studies to elucidate the mechanisms and removal capacity of ARB and ARGs with combined UV and chlorination. Inactivation of ARB and ARGs through UV disinfection with chlorine is shown in

Table 6. ARB and ARGs inactivation through UV/Cl₂ disinfection.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Secondary treated wastewater	40 mJ/cm2 Cl2: 2 mg/L t = 10 min	aacC2, sulII, strB, tetA and tetB	0.41–0.94	[91]
	32 mJ/cm2 Cl2: 1–2 mg/L t = 20–80 s	antibiotic resistant Morganella Morganii and Enterococcus Faecalis	1.4 log
Secondary treated effluent	62.5 mJ/cm2 Chlorine: 15 mg/L t = 30 min	sulI, tetX, tetG	0.30–0.50 log	[98]
	124.8 mJ/cm2 Chlorine: 30 mg/L t = 30 min		1–1.4 log
	249.5 mJ/cm2 Chlorine: 30 mg/L t = 30 min		1.7–2.1 log

.

5.1.3. UV/Persulfate

UV could activate persulfate anion (PS, S₂O₈²⁻) and sulfate radicals (SO₄·⁻) are produced [158,159]. Due to the reaction of SO₄·⁻ with water, HO·, are formed as a function of the pH. However, sulfate radicals are the major activated radical species [159,160]. Sulfate radicals (E^o = 2.5–3.1 V, half-life = 30–40 μS) have a stronger redox potential, longer half-life time and a flexibility over a broad pH range compared to hydroxyl radicals (E^o = 1.89–2.72 V, half-life = 10⁻³ μS), rendering them a promising tool for the removal of pollutants and bacteria from wastewater (Table 7) [137,161]. SO₄·⁻ can cause detrimental effects to the membrane permeability, to the intracellular components of bacteria and to cellular reproduction, while UV injure DNA, causing the inactivation of bacteria [162].

Many studies found that the UV/PS exhibited a higher and more efficient removal of ARB and ARGs than UV alone [127,163]. Moreover, they attributed the reduction of ARGs and ARB to the high removal of mobile genetic elements (76.09%), which are considered as drivers of the spread of antibiotic resistance. Serna-Galvis et al. [164] showed that UV/PS and UV-C/H₂O₂ could remove more than 98% of bla_KPC-3 in only 5 min. However, the initiative inactivation upon 60 s of treatment was higher in the UV/PS system possibly due to the longer half-time of sulfate radicals and due to the higher rate of strand breakage by sulfate radicals. Michael-Kordatou et al. [159] observed the complete removal of erythromycin-resistant E. coli from secondary treated wastewater effluent by UV-C/PS. The required time to inactivate E. coli by UV-C/PS was 45 min while the inactivation by UV-C alone was 90 min, showing the enhancement due to the presence of PS. However, more research is needed to fill in the gaps in knowledge regarding the effectiveness of UV/PS in inactivating ARB and reducing the amount of ARGs.

Table 7. Inactivation of ARB and ARGs by .UV/PS disinfection.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Secondary treated effluent	1 mmol/L PS UV dose: 240 mJ/cm2	macrolides, sulfonamides, tetracyclines, quinolones-resistant bacteria	96.6, 94.7, 98, 99.9%	[127]
		sulI, sulII, ermB, qnrS and tetO	3.84 log
Secondary treated effluent	P: 8W, [PS]: 1.0 mM 180 s	carbapenem resistant Klebsiella pneumoniae	6 log	[164]
	P: 8W, [PS]: 1.0 mM 5 min	blaKPC-3	>98%
Secondary effluent	UV-C/ 10 mg/L PS 45 min,	erythromycin-resistant E. coli	100%	[159]
Tertiary treated urban wastewater effluent	[PS]: 2mM pH = 6.9–7.0 UV dose: 216 mJ/cm2	multi-resistant E. coli	6.7 log	[163]
		aphA, tetA	>3 log

Table 7. Inactivation of ARB and ARGs by .UV/PS disinfection.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Secondary treated effluent	1 mmol/L PS UV dose: 240 mJ/cm2	macrolides, sulfonamides, tetracyclines, quinolones-resistant bacteria	96.6, 94.7, 98, 99.9%	[127]
		sulI, sulII, ermB, qnrS and tetO	3.84 log
Secondary treated effluent	P: 8W, [PS]: 1.0 mM 180 s	carbapenem resistant Klebsiella pneumoniae	6 log	[164]
	P: 8W, [PS]: 1.0 mM 5 min	blaKPC-3	>98%
Secondary effluent	UV-C/ 10 mg/L PS 45 min,	erythromycin-resistant E. coli	100%	[159]
Tertiary treated urban wastewater effluent	[PS]: 2mM pH = 6.9–7.0 UV dose: 216 mJ/cm2	multi-resistant E. coli	6.7 log	[163]
		aphA, tetA	>3 log

5.2. Photo-Fenton Process

Fenton process is the catalysis of hydrogen peroxide by ferrous iron that generates reactive hydroxyl radicals (HO·) under acidic conditions [165]. The light source enhances the formation of reactive oxidizing species about 40 times than the dark Fenton, thus providing a greater potential for the inactivation of intracellular DNA [166,167]. Natural sunlight has the ability to inactivate pathogens and ARB due to UV wavelengths, damage DNA, and thus further increase the disinfection potential of solar photo-Fenton treatment [168,169,170,171]. The implementation of solar irradiation is a sustainable and cost-effective method because it decreases the energy demand [172]. Solar photo-Fenton treatment produces photochemical active complexes with iron due to the presence of natural organic matter in wastewater effluents, accelerating the production of hydroxyl radicals, enhancing the disinfection potential [173]. The removal of ARB and ARGs by Fenton/photo-Fenton oxidation process depends on the concentration of H₂O₂, Fe²⁺, pH, and time of exposure as shown in Table 8. Another key point is that solar-photo Fenton can efficiently remove the total DNA from wastewater and, therefore, hampering the spread of antibiotic resistance to other organisms in the environment [102]. The iron precipitates in neutral pH, and it has been found that the optimum pH for the efficient implementation of photo Fenton is 2.8 [174]. In addition, the optimum pH for the inactivation of ARGs was found to be three [175], but it is not known if the high removal is due to the high formation of hydroxyl radicals or to the effect of acidic conditions to the viability of bacteria [102]. Michael et al. [165] noticed a 1 log reduction of ARB after the acidification of wastewater. Moreover, numerous studies have reported the efficient removal of ARB at an acidic pH [165,173]. The acidic pH requirement makes full-scale application of the method difficult, as the pH of secondary treated wastewater is neutral, and there is a need of acidification and subsequent neutralization before disposal or reuse, elevating the operating costs [176]. However, various studies have proved the efficiency of solar photo-Fenton to remove ARB and ARGs at neutral pH. Ioannou-Ttofa et al. [177] examined the efficiency of solar photo-Fenton at pH 8 and total inactivation of ampicillin-resistant E. coli occurred after 120 min of treatment. Moreover, no regrowth of ampicillin-resistant bacteria was observed after 48 h of incubation after the application of solar-Fenton, indicating permanent damage of bacterial cells. These observations were consistent with Michael et al. [165] when the solar photo-Fenton was applied for the inactivation of fecal coliforms, Enterococcus spp., P. aeruginosa, and total heterotrophs resistant to erythromycin, ofloxacin, and trimethoprim under acidic pH, and in addition, no re-activation had arisen after 24 h of treatment at 25°C. Solar photo-Fenton seems to be an ideal method to permanently inactivate the metabolic activity and growth of bacteria.

In an attempt to test the effect of solar photo-Fenton at neutral pH, Fiorentino et al. [97] used nitrilotriacetic acid (NTA) as a chelating agent to prevent iron precipitation for the removal of intI, sulI, bla_TEM, qnrS, tetM from secondary treated wastewater. They found that under the conditions of 2 mM Fe³⁺-NTA with 4.41 mM H₂O₂, the removal of target genes was not possible. However, they also found that under acidic conditions there was only a small reduction of 23% and 26% for intI and sulI, respectively. On the contrary, Vilela et al. [167] supplied Fe²⁺ intermittently in order to provide a high quantity of Fe²⁺ at a neutral pH. This method was effective at removing 80% of plasmids conferring resistance to ampicillin and kanamycin from secondary municipal wastewater. Furthermore, the removal efficiency of these species by dark Fenton was 53%, showing the enhancement due to the irradiation. Ahmed et al. [143] tested a novel material of nanopyrite (FeS₂) on graphene oxide (FeS₂@ GO) for the inactivation of ARB and ARGs from wastewater under simulated solar irradiation at a neutral pH. They observed that a dose of 0.25 mg/L FeS₂@ GO catalyst and 1.0 mM H₂O₂ was able to remove ARB by 6 log in 30 min of exposure and the extracellular ARGs by 7 log in only 7.5 min of exposure from synthetic wastewater. However, the intracellular ARGs were removed by 1.20 log in 30 min of treatment while an exposure time of 2 h was required for a removal of 4 log. The removal of ARB in real wastewater was lower (4.43 log) due to the presence of ROS scavengers, such as bicarbonates. They attributed the inactivation of ARB and ARGs to the generation of reactive oxygen species, such as hydroxyl radicals (HO·), superoxide radicals (O₂·⁻), and singlet oxygen (O₂).

Table 8. Inactivation of ARB and ARGs by photo Fenton.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Secondary treated wastewater	0.1 mM of Fe2+ with 1.47 mM of H2O2, acidic pH, natural sunlight	intI and sulI	23% and 26%	[97]
	0.2 mM of Fe3+-NTA with 4.41 mM of H2O2, neutral pH, natural sunlight	intI, sulI, blaTEM, qnrS, tetM	No removal
Secondary treated wastewater	50 mg L of H2O2 and 30 mg/L of Fe2+268 w/m2 (solar simulator) 60 min neutral pH	Plasmids conferring resistance to ampicillin and kanamycin	80%	[167]
Synthetic secondary wastewater	50 mg/L of H2O2 and 30 mg/L of Fe2+ 268 w/m2 (solar simulator) 60 min neutral pH	Plasmids conferring resistance to ampicillin and kanamycin	100%	[167]
		penicillin-, cefalexin-resistant bacteria	100%
Synthetic wastewater	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 30 min	E. coli resistant to tetracycline and ampicillin	6 log	[143]
	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 7.5 min	e-blaTEM-1,e-tetA	7 log
Real wastewater effluent	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 30 min	E. coli resistant to tetracycline and ampicillin	4.43 log	[143]
Synthetic wastewater	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 2 h	i-blaTEM-1, i-tetA	4 log	[143]
Secondary treated wastewater	[Fe2+]0 = 5 mg/L, [H2O2]0 = 75 mg/L, pH: 8.0, 63 W/m2 (solar simulator)	Ampicillin-resistant E. coli	Complete inactivation	[177]
Conventional activated sludge	[Fe2+]0 = 5 mg/L, [H2O2]0 = 100 mg/L, pH = 2.8–2.9 natural sunlight 180 min	fecal coliforms, Enterococcus spp., P. aeruginosa, and total heterotrophs resistant to eryth-romycin, ofloxa-cin, trimethoprim	Complete inactivation	[165]
Conventional activated sludge	[Fe2+]0 = 5 mg/L, [H2O2]0 = 100 mg/L, pH= 2.8–2.9 natural sunlight 180 min	blaOXA, blaCTX-M, qnrS, sulI, tetM	<LOD	[165]
Secondary treated effluent	Fe3+ = 5 mg/L, H2O2 = 50 mg/L, pH = 4 natural sunlight 120 min	clarithromycin and sulfamethox-azole-resistant Enterococcus	5 log	[173]
MBR effluent	[Fe2+]0 = 5 mg/L, [H2O2]0 = 50 mg/L, pH = 2.8 63 W/m2 (solar simulator)	mecA, ermB, sulI, ampC	<LOD	[52]

Table 8. Inactivation of ARB and ARGs by photo Fenton.

Treatment	Conditions	ARB/ARGs	Removal Efficiency	Reference
Secondary treated wastewater	0.1 mM of Fe2+ with 1.47 mM of H2O2, acidic pH, natural sunlight	intI and sulI	23% and 26%	[97]
	0.2 mM of Fe3+-NTA with 4.41 mM of H2O2, neutral pH, natural sunlight	intI, sulI, blaTEM, qnrS, tetM	No removal
Secondary treated wastewater	50 mg L of H2O2 and 30 mg/L of Fe2+268 w/m2 (solar simulator) 60 min neutral pH	Plasmids conferring resistance to ampicillin and kanamycin	80%	[167]
Synthetic secondary wastewater	50 mg/L of H2O2 and 30 mg/L of Fe2+ 268 w/m2 (solar simulator) 60 min neutral pH	Plasmids conferring resistance to ampicillin and kanamycin	100%	[167]
		penicillin-, cefalexin-resistant bacteria	100%
Synthetic wastewater	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 30 min	E. coli resistant to tetracycline and ampicillin	6 log	[143]
	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 7.5 min	e-blaTEM-1,e-tetA	7 log
Real wastewater effluent	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 30 min	E. coli resistant to tetracycline and ampicillin	4.43 log	[143]
Synthetic wastewater	0.25 mg/L FeS2@ GO 1 mM H2O2 130 mW/cm2 (solar simulator) 2 h	i-blaTEM-1, i-tetA	4 log	[143]
Secondary treated wastewater	[Fe2+]0 = 5 mg/L, [H2O2]0 = 75 mg/L, pH: 8.0, 63 W/m2 (solar simulator)	Ampicillin-resistant E. coli	Complete inactivation	[177]
Conventional activated sludge	[Fe2+]0 = 5 mg/L, [H2O2]0 = 100 mg/L, pH = 2.8–2.9 natural sunlight 180 min	fecal coliforms, Enterococcus spp., P. aeruginosa, and total heterotrophs resistant to eryth-romycin, ofloxa-cin, trimethoprim	Complete inactivation	[165]
Conventional activated sludge	[Fe2+]0 = 5 mg/L, [H2O2]0 = 100 mg/L, pH= 2.8–2.9 natural sunlight 180 min	blaOXA, blaCTX-M, qnrS, sulI, tetM	<LOD	[165]
Secondary treated effluent	Fe3+ = 5 mg/L, H2O2 = 50 mg/L, pH = 4 natural sunlight 120 min	clarithromycin and sulfamethox-azole-resistant Enterococcus	5 log	[173]
MBR effluent	[Fe2+]0 = 5 mg/L, [H2O2]0 = 50 mg/L, pH = 2.8 63 W/m2 (solar simulator)	mecA, ermB, sulI, ampC	<LOD	[52]

6. Conclusions

In this review, we evaluated the potential removal of ARB and ARGs with disinfection methods, advanced oxidation processes, and tertiary wastewater treatment. Conventional wastewater treatment plants are not capable of efficiently removing ARB and ARGs, and therefore, their coupling with advanced wastewater treatment processes is essential to hinder the spread of antibiotic resistance in water bodies. The revision of the current Urban Wastewater Treatment Directive 91/271/EEC requires wastewater treatment plants to upgrade their facilities with advanced wastewater treatment processes to remove micropollutants. Based on the scientific literature on the removal of antibiotic-resistant bacteria (ARB) and antibiotic resistance genes (ARGs) with disinfection methods, advanced oxidation processes, and tertiary wastewater treatment, the main findings can be summarized as follows:

• The inactivation of ARB and ARGs is influenced by the type and operating parameters of the treatment. ARGs are more recalcitrant to treatments, which poses the risk of their dissemination and prevalence in environmental microorganisms;
• The removal of ARGs with conventional disinfection methods is not clearly understood. For example, the removal of a significant portion of ARGs requires a higher UV dose than is normally used in wastewater treatment plants. Therefore, further experimental studies are needed to investigate the removal of antibiotic resistance with combined technologies, such as UV/chlorine and UV/persulfate. The presence of oxidative radicals enhances the inactivation of ARB and ARGs;
• Further studies are needed to clarify the reactivation potential of ARB after UV-induced AOPs, when the oxidative stress is relieved;
• Solar photo-Fenton appears to be a promising treatment for the removal of ARB and ARGs from wastewater. It is also an environmentally friendly method because it uses solar light, making it a promising tool for countries with abundant sunlight. However, more research is needed on the removal potential at neutral pH, which is prevalent in secondary treated wastewater effluent;
• The studies reviewed focused only on a limited number of ARG types quantified with quantitative polymerase chain reaction (qPCR), which generally confer resistance to tetracyclines, sulphonamides, and β-lactams. However, there are thousands of ARGs that may be present in wastewater. Metagenomic sequencing could be a powerful tool to identify a wide range of ARGs;
• Sand and granular activated carbon filters can be used to remove disinfection by-products and may also improve the removal of ARB and ARGs, but further studies are needed;
• The development of new materials, such as nanomaterials and adsorbents, for the removal of ARGs and ARB from wastewater is of interest. The removal potential and mechanisms of these materials should be investigated;
• There is an urgent need to clarify the impact of antibiotic resistance present in the environment on human health. Epidemiological studies should be conducted to determine whether ARB and ARGs present in bathing waters, where treated wastewater is discharged, as well as in crops irrigated with treated wastewater, could lead to their colonization in the human body;
• As water scarcity is becoming a major global challenge, wastewater reuse can be an alternative source of water for irrigation and industrial use. However, further research is needed on how advanced wastewater treatments can prevent the presence of ARB and ARGs in treated wastewater to protect public health and the environment;
• Future research should address mitigation strategies related to wastewater reuse and sustainability. For example, adsorbents (e.g., activated carbon) can be used as a cost-effective and environmentally friendly approach.