Antibiotic resistant bacteria (ARB) with antibiotic resistance genes (ARGs) can reduce or eliminate the effectiveness of antibiotics and thus threaten human health. The United Nations Environment Programme considers antibiotic resistance the first of six emerging issues of concern. Advanced oxidation processes (AOPs) that combine ultraviolet (UV) irradiation and chemical oxidation (primarily chlorine, hydrogen peroxide, and persulfate) have attracted increasing interest as advanced water and wastewater treatment technologies. These integrated technologies have been reported to significantly elevate the efficiencies of ARB inactivation and ARG degradation compared with direct UV irradiation or chemical oxidation alone due to the generation of multiple reactive species. In this study, the performance and underlying mechanisms of UV/chlorine, UV/hydrogen peroxide, and UV/persulfate processes for controlling ARB and ARGs were reviewed based on recent studies. Factors affecting the process-specific efficiency in controlling ARB and ARGs were discussed, including biotic factors, oxidant dose, UV fluence, pH, and water matrix properties. In addition, the cost-effectiveness of the UV-based AOPs was evaluated using the concept of electrical energy per order. The UV/chlorine process exhibited a higher efficiency with lower energy consumption than other UV-based AOPs in the wastewater matrix, indicating its potential for ARB inactivation and ARG degradation in wastewater treatment. Further studies are required to address the trade-off between toxic byproduct formation and the energy efficiency of the UV/chlorine process in real wastewater to facilitate its optimization and application in the control of ARB and ARGs.
Antibiotics are considered one of the most momentous inventions in pharmaceutical science because antibiotics effectively treat bacterial infections and increase life expectancy [1]. Globally, antibiotic consumption has been estimated at 100 000-200 000 tonnes per year [2], and the use of antibiotics continues to rise. A survey of antibiotic consumption in 76 countries revealed that antibiotic consumption increased by 65% from 2000 to 2015 [3]. Antibiotic resistance occurs naturally when bacteria evolve to become resistant to the effects of antibiotics [4]. However, anthropogenic activities, including misuse or overuse of antibiotics, greatly promote the selection of antibiotic resistant bacteria (ARB) and antibiotic resistance genes (ARGs) [5], [6]. Antibiotic resistance threatens the efficacy of antibiotics in treating infectious diseases, resulting in longer hospital stays, higher therapeutic costs, and higher morbidity and mortality rates [5], [7], [8]. It was predicted that ten million people would die annually by 2050 as a result of antimicrobial resistance [9]. The World Health Organization has declared that antibiotic resistance is one of the most severe threats to public health in the 21st century [9].
Intrinsic resistance, genetic mutations, vertical gene transfer, and horizontal gene transfer have been reported as primary mechanisms for the development and spread of antibiotic resistance [10], [11]. Intrinsic resistance is related to actions such as the prevention of antibiotic entry into the cytoderm, expulsion of specific antibiotics by explicit efflux pumps before they reach their targets, and degradation or even deactivation of antibiotics. Mutagenesis leads to antibiotic resistance in several ways, including switching antibiotic targets, medically altering enzymes, increasing drug efflux, and decreasing target gene expression [12]. Vertical gene transfer means that the genetic material is inherited from parental cells and occurs only when microbial gene portability is possible for transfer to the daughter cells [13]. Although the level of extracellular ARGs (e-ARGs) is typically 2-3 orders of magnitude lower than that of intracellular ARGs (i-ARGs), the spread of e-ARGs could be significantly enhanced by horizontal gene transfer [10], [14]. Horizontal gene transfer can occur through gene transfer agents, conjugation, transformation, and transduction (Fig. 1[10,15]). Gene transfer agents are virus-like elements that transfer parts of DNA from host cells to recipient cells or free particles that spread to recipient cells by cell lysis [10]. Conjugation occurs when cell-to-cell contact causes ARGs carried by mobile genetic elements to be delivered from donor to recipient cells [16]. In the transformation process, naturally transformable bacteria uptake, integrate, and functionally express e-ARGs. In the transduction process, i-ARGs from a donor cell are packaged into bacteriophages when DNA replication begins, and ARGs carried by bacteriophages are combined with the genome of recipient cells once they are infected by bacteriophages [17]. The multiple mechanisms for the emergence and dissemination of antibiotic resistance highlight the challenges in controlling ARB and ARGs [18], [15].
It has been reported that approximately 50%-90% of human and veterinary antibiotics and their metabolites are excreted in their active forms via urine and feces [19]. Wastewater containing antibiotics from households, livestock farms, and hospitals continuously enters wastewater treatment plants [20], where most antibiotics are inefficiently removed or deactivated by traditional wastewater treatment processes [21], [22]. At the same time, activated sludge used in the biological treatment unit has high microbial density and diversity, which favors horizontal gene transfer [23]. With relatively high levels of antibiotics and substantial amounts of activated sludge biomass, wastewater treatment plants have been reported to be hotspots for promoting the proliferation of ARB and ARGs [24], [25]. For instance, the relative abundance of tet genes in wastewater effluents increased by 212%-358% after biological treatment [24]. Due to the insufficient removal efficiency of conventional wastewater treatment, significant amounts of antibiotics, ARB, and ARGs may enter aquatic environments [25], [26]. As a result, the occurrence of ARB and ARGs from wastewater to drinking water sources has been frequently detected [24], [25], [26], [27], [28], [29]. Recently, Zhang et al. [4] collected 4572 metagenomic samples from the European Nucleotide Archive to determine the global distribution of ARGs, among which 1819 samples were from wastewater and aquatic environments. A total of 2561 ARGs that collectively conferred resistance to 24 classes of antibiotics were identified. Only 25 ARGs were detected with an occurrence frequency of higher than 75%, and the frequency of most ARGs was less than 10%. The abundance of ARGs in aquatic environments varied markedly, ranging from less than 0.005 to 296 reads per kilobase per million per sample. By considering the occurrence frequency and abundance of ARGs and other factors such as the potential of contributing to pathogenicity, ARGs with relatively high health risks were determined and are listed in Table 1 [4].
Ultraviolet (UV) and chemical oxidants (e.g., chlorine) are commonly used in water and wastewater treatment for disinfection, which is considered the last and most solid barrier to control microbial risks [30]. To inactivate ARGs, higher UV doses are required than those used in standard disinfection practices [26], [31]. However, there is evidence that UV and chemical oxidation might enrich ARB and ARGs in finished water/wastewater [28], [32], [33], [34], [35]. For instance, drinking water chlorination was found to increase the absolute abundance of 14 of the 18 ARG subtypes [36]. In a full-scale wastewater treatment plant, chlorination was found to increase the absolute abundance of e-ARGs by up to 3.8 times and i-ARGs by 7.8 times [33]. Yuan et al. [34] reported that chlorination increased the transformation frequency of e-ARGs by 2.9-7.2 times because chlorination resulted in cell debris with enhanced adsorption of e-ARGs and thus increased their persistence and propagation potential. UV treatment was found to increase the relative abundance of 159 ARGs in wastewater by 6.0 times [32]. In addition, ARB may regenerate and reactivate in water/wastewater treated with UV or chemical oxidants [37], [38]. Such regrowth of damaged cells may lead to horizontal gene transfer and the spread of antibiotic resistance [39], further compromising the microbial safety of water/wastewater. These findings indicate that treatment by UV irradiation or chemical oxidation alone is not adequate for effectively inactivating ARB and degrading ARGs.
In recent years, controlling the risks of ARB and ARGs in aquatic environments with UV-based advanced oxidation processes (AOPs) has gained increasing interest, as these processes have been reported to significantly promote the reduction of ARB and ARGs due to the rapid formation of reactive species [40], [41], [42], [43], [44], [45], [46], [47], [48]. The application and performance of UV-based advanced treatments in controlling ARB and ARGs have been previously reviewed [31], [49]. In this study, a systematic literature review of the state of the art of ARB/ARG control by UV-based AOPs was conducted. UV-based AOPs mainly include UV/hydrogen peroxide (H2O2), UV/chlorine, UV/persulfate, UV/ozone, and UV/catalyst processes. Few studies have been reported on ARB/ARG treatment with UV/ozone and UV/catalyst processes [43], [50], [51], [52], [53], and therefore this study focused on UV/H2O2, UV/chlorine, and UV/persulfate processes. The efficiencies and mechanisms of the three UV-based AOPs in inactivating ARB and degrading ARGs under various environmental and operational conditions were discussed, with emphasis on improving the fundamental understanding of ARB and ARG control and providing guidelines for the engineering design of UV-based AOPs. Moreover, the cost-effectiveness of the three UV-based AOPs was systematically evaluated and compared based on the concept of electrical energy per order (EE/O), which may provide decision-makers with useful insights for selecting appropriate UV-based AOPs for practical application.
2. Inactivation of ARB and degradation of ARGs by UV-based AOPs
2.1. Efficiencies and mechanisms of UV-based AOPs in ARB inactivation and ARG degradation
The UV/chlorine process is gaining interest as an AOP due to its ease of implementation and high efficiency in degrading recalcitrant micropollutants, including ARB and ARGs [54], [55], [56]. Table 2 [44], [56], [57], [58], [59], [60] shows data for ARB inactivation and ARG reduction in the combined UV/chlorine, UV alone, and chlorination alone. UV/chlorine has shown a robust capability of inactivating different kinds of ARB, with higher efficiencies than the summed efficiencies of UV alone and chlorination alone [49], [57]. Synergistic effects were also found for ARG degradation using the UV/chlorine treatment. Compared with chlorine (20 mg·L−1) or UV254 (9.03 mW·cm−2) alone, 40 min of UV/chlorine treatment increased the degradation of tetM and blaTEM by 0.98-3.20 log and 1.28-3.36 log, respectively, in phosphate-buffered saline (PBS) [57]. In wastewater treatment plants, the combination of UV254 (0.1 mW·cm−2) and chlorination (2 mg·L−1) achieved an additional 1.4 log ARB in 1.3 min and 1.0-1.5 log ARG reduction in 53 min [60]. Furthermore, regrowth and reactivation of ARB were significantly inhibited by UV/chlorine treatment.
The UV/H2O2 process has also been extensively investigated to mitigate the risks associated with ARB and ARGs in aquatic environments [61], [62], [63]. In a 30 min UV254 (9.85 mW·cm−2)/H2O2 (340 mg·L−1) treatment, the abundances of target ARGs (sul1, tetX, tetG, and intI1) in wastewater were significantly reduced by 1.55-2.32 log [64]. Michael et al. [65] also reported substantial reductions of target ARGs (i.e., 2.0-3.7 log reduction of sul1, sul2, tetM, blaOXA-A, and blaTEM) by a 90 min UV254 (3.46 mW·cm−2)/H2O2 (5 mg·L−1) treatment. In contrast, an insignificant reduction in the abundances of target ARGs (including blaTEM, qnrS, and tetW) was observed by up to 240 min UV320-450 (17.4 μW·cm−2)/H2O2 (20 mg·L−1) treatment of wastewater [66]. As shown in Table 3 [41], [45], [62], [63], [64], [65], [66], [67], [68], [69], relatively high UV fluence and chemical doses were required for effective ARB inactivation and ARG degradation by UV/H2O2 treatment, while the efficiency was still higher than that of the H2O2 alone or UV alone process. The UV/H2O2 process has been shown to significantly inhibit the photoactivation and regeneration of bacteria [70].
Activation of persulfate, including peroxydisulfate (PDS) and peroxymonosulfate (PMS), by UV irradiation has shown promising potential to combat ARB and ARGs [45], [46], [47], [71]. For instance, the UV254 (2.3 μW·cm−2)/PDS (238 mg·L−1) process was more effective than the UV254 (2.3 μW·cm−2)/H2O2 (34 mg·L−1) process in removing blaKPC-3 in wastewater, and the two processes resulted in 80% and 67% reductions in 1 min, respectively [45]. For a 5 min treatment, both UV254/PDS and UV254/H2O2 treatments achieved higher than 98% degradation of the blaKPC-3 gene. UV254 (100 μW·cm−2)/PMS (20 mg·L−1) treatment was found to reduce sul1 and intI1 levels by 2.9 log and 3.4 log, respectively, in 30 min [42]. The treatment efficiencies of ARB and ARGs using UV254/PDS and UV254/PMS under various operational conditions are listed in Table 4 [42], [45], [46], [47], [48], [71], [72], [73]. The combination of UV254 and persulfate has been reported to effectively inactivate different types of ARB (e.g., antibiotic-resistant Escherichia coli (E. coli), Klebsiella pneumoniae (K. pneumonia), and Pseudomonas sp. HLS-6), and synergistic effects were observed compared with individual UV or persulfate treatment [45], [46], [47]. However, the superiority of the combined process over the individual UV254 or persulfate process varied for different ARGs. Zhou et al. [46] observed that UV254 (0.4 μW·cm−2)/PDS (238 mg·L−1) treatment achieved a higher total ARG reduction level (3.84 log) than the UV254 irradiation (3.28 log) and PDS oxidation (1.68 log). Specifically, 83.4% of macrolide resistance genes were degraded by UV254/PDS, which was significantly higher than the 68.5% reduction by UV254 irradiation and the 35.9% reduction by PDS oxidation. In contrast, the combined UV/PDS process improved the degradation of quinolone resistance genes by less than 5% compared with UV irradiation alone [46].
Compared to UV irradiation or chemical oxidation alone, the generally higher reduction of ARB and ARGs achieved by the UV-based AOPs has been reported to relate with the enhanced destruction of the spore surface and the distinct release of intracellular materials from ARB [61], [72]. Effective damage to cell membrane integrity by UV-based AOPs promoted the inactivation of ARB and inhibited their regrowth. Fig. 2 [40,74] shows a schematic illustration of how ARB and ARGs are removed by UV-based AOPs. Although UV light can damage cell wall macromolecules, the main mechanism of bacterial inactivation by UV irradiation is to cause DNA double-strand breakage and chromosome aberration, leading to cell death by necrosis or apoptosis [75]. Chemical oxidants have been reported to inactivate bacteria primarily by damaging the cell surface and altering membrane permeability [76], with chlorine having a higher disinfection potency than H2O2 and persulfate [77], [78]. During the UV/chlorine, UV/H2O2, or UV/persulfate process, multiple radicals can be generated (Fig. 2, detailed in Section 2.2). The generated radicals are highly oxidative and can cause pronounced cell surface damage and increased cell permeability [79], [74], which may facilitate the reaction of oxidants and free radicals with ARG-encoding DNA and promote DNA damage by UV irradiation [13]. Meanwhile, the increased cell permeability allows i-ARGs to be released into aquatic environments and converted into e-ARGs. It has been shown that compared with i-ARGs, e-ARGs are more likely to be degraded because they are more readily exposed to UV irradiation, chemical oxidants, and free radicals in the UV-based AOPs [74], [80], [81].
2.2. Formation and roles of reactive species
As described in Eqs. (1)-(6), multiple active oxidants are produced in the UV/chlorine process, including hydroxyl radicals (∙OH) and reactive chlorine species (RCS) such as Cl∙, Cl2∙−, and ClO∙ (Fig. 2) [31]. The contributions of different radicals in the treatment of ARB and ARGs have been studied by using radical scavengers such as nitrobenzene (NB) for ∙OH scavenging (the reaction rate constant k∙OH-NB=3.9×109L·mol-1·s-1) in the UV/chlorine process. The addition of nitrobenzene resulted in little change in the log reduction of ARGs, indicating that ∙OH insignificantly contributed to ARG degradation in the UV/chlorine process. Similar results were obtained in previous studies [41], [82], which indicated that the minor contribution of ∙OH to ARG degradation was because ∙OH could be easily consumed by cell surface components and the intracellular matrix. However, in a sample with benzoic acid as the scavenger of both ∙OH and RCS, the reduction of tetM and blaTEM abundances was comparable to that with nitrobenzene as a scavenger of ∙OH only [57], indicating that ∙OH instead of RCS primarily accounted for the tetM and blaTEM degradation in the UV/chlorine process. The inconsistency in the literature regarding the role of various radicals in ARG degradation could be ascribed to differences in bacterial strains and experimental conditions (e.g., UV and chlorine doses).
$\mathrm{HOCl} / \mathrm{OCl}^{-}+h v \rightarrow \mathrm{HO}^{·} / \mathrm{O}^{·-}+\mathrm{Cl}^{·}$
In the UV/H2O2 process, ∙OH is generated via the photolysis of H2O2 during the UV/H2O2 process, while a high concentration of H2O2 may serve as a ∙OH scavenger, as shown in Eqs. (7)-(10) [67], [83]. Due to the high reactivity and non-selectivity of ∙OH, it is rapidly consumed by reactions with cell surface components, leading to its negligible impact on i-ARGs [41], [63]. However, Meng et al. [48] observed a much greater degradation of i-ARGs with the UV/H2O2 process than with UV treatment alone. This might be because the permeability of the bacterial membrane can be altered by ∙OH with increasing treatment time, allowing the penetration of UV irradiation and H2O2 into the cell and subsequent interactions with intracellular ARG-encoding DNA. In addition, it has been reported that the cytosol of bacteria can react with H2O2 to form ∙OH, which can reduce enzyme activities, weaken metabolism, and cause mutagenic structural damage to DNA and RNA [45], [84].
$\mathrm{H}_{2} \mathrm{O}_{2}+h v \rightarrow 2 \mathrm{HO}^{·}$
The UV/PDS and UV/PMS processes are new AOPs that primarily produce ∙OH and SO4∙− (Fig. 2). The main chemical reactions involved are shown in Eqs. (11), (12), (13), (14), (15) at near neutral pH [85], [86]. The quantum yields (Φ) of radicals from UV photolysis of PDS at 254 nm in deoxygenated (Φ = 1.4) and oxygen-saturated (Φ = 1.8) water are significantly higher than those of PMS (Φ = 1.04) and H2O2 (Φ = 1.0) [87]. Compared with ∙OH, SO4∙− has higher selectivity, a longer half-life (30-40 μs for SO4∙− and 20 ns for ∙OH), and a higher redox potential (2.5-3.1 V for SO4∙− and 2.8 V for ∙OH) [88]. Moreover, SO4∙− can attack the guanine heterocycle at the π-site, resulting in DNA modification [42]. Accordingly, SO4∙− has been reported to be the predominant reactive species responsible for inactivating ARB and degrading ARGs in the UV/PDS and UV/PMS processes [45]. However, the role of different radicals generated by UV/PDS or UV/PMS in bacterial inactivation and gene degradation may be influenced by bacterial strains and gene types. For instance, when tetracycline resistance genes in marine agricultural wastewater were treated with UV/PDS, SO4∙− and ∙OH mainly accounted for the degradation of tetA and tetW, respectively, and both radicals contributed significantly to the degradation of tetM [89].
$\mathrm{S}_{2} \mathrm{O}_{8}{ }^{2-}+h v \rightarrow 2 \mathrm{SO}_{4}{ }^{·-}$
$\mathrm{HSO}_{5}^{-}+h v \rightarrow \mathrm{SO}_{4}^{·-}+\mathrm{HO}^{·}$
3. Factors influencing the efficiencies of the UV-based AOPs
3.1. Biotic factors
Extracellular polymeric substances (EPS) are organic polymers of microbial origin and include polysaccharides, proteins, extracellular DNA, and lipids [90]. As a fundamental component of biofilms, EPS can enhance the resistance of bacteria to environmental stress. EPS acts as a shield that protects microbes from external damage, such as liquid shear force, antibiotics, and UV irradiation [48], [91]. For UV-based AOPs, the dense EPS layer should be degraded first (e.g., by highly reactive radicals through oxidation of polysaccharides and proteins in EPS), and then UV irradiation, chemical oxidants, and free radicals can interact with the bacterial cells. Therefore, the treatment efficiencies of UV-based AOPs for ARB and ARGs are significantly affected by the EPS matrix. The treatment efficiencies are also related to bacterial type, as different bacteria have different resistances to external perturbation or destruction by UV, oxidants, and radicals. Jia et al. [92] observed that chlorination effectively reduced the levels of Methylophilus, Methylotenera, Limnobacter, and Polynucleobacter, while chlorination increased the relative abundance of Pseudomonas, Acidovorax, Sphingomonas, Pleomonas, and Undibacterium in drinking water. As UV fluence increased, the relative abundance of Gram-positive ARB in the microbial community increased [93]. This was attributed to the thicker peptidoglycan layers of Gram-positive ARB, which can inhibit UV penetration [93]. In addition, the smaller total genome size of Gram-positive ARB may reduce their susceptibility to UV irradiation because they have fewer latent pyrimidine dimers, the prominent targets of UV irradiation in DNA [94]. The initial levels of ARB could also impact treatment performance. A lower initial concentration of ARB may be associated with a higher inactivation efficiency [42] because the bacterial cytoplasm could release high concentrations of organics that consume radicals during UV-based AOP treatment. For the destruction of ARGs, the properties and structures of ARGs are crucial influencing factors. It has been reported that the efficiencies of UV and oxidants in degrading different types of genes are related to the target site, guanine content, and the number of latent dimers [42], [95]. For instance, UV-induced ARG degradation was proportional to the number of adjacent thymine sites because UV irradiation causes DNA damage mainly through thymine dimerization [94].
3.2. Oxidant dose and UV fluence
Compared with ARB, the removal of ARGs by oxidants or UV irradiation usually requires much higher doses of UV and oxidants than those in conventional processes. However, the degradation of ARGs is not always proportional to the oxidant dose, as the overdosing of oxidants may cause a self-scavenging effect in radicals and an inner filter effect by competing for photons with target ARGs [38], [42]. The optimal oxidant dose may vary for different target genes [44]. For instance, in the UV/chlorine process, the degradation of sul1 gene increased as the chlorine dose was increased up to 5 mg·L−1, while no further enhancement was observed at higher chlorine doses; in contrast, the optimal chlorine dose for intI1 gene removal was 20 mg·L−1 [44]. Zhang et al. [54] also reported that the synergism of ARG degradation increased significantly with chlorine dose from 15 to 25 mg·L−1 in the UV/chlorine process, whereas continuous increases in the chlorine dose resulted in an insignificant difference in ARG reduction. In the UV/H2O2 process, the maximum reduction of several commonly detected ARGs (e.g., sul1, intI1, tetX, and tetG) was achieved at an H2O2 dose of 340 mg·L−1 [64]; further increases in the H2O2 dose resulted in lower degradation efficiencies for the target ARGs, which was attributed to ∙OH scavenging by higher concentrations of H2O2. Similarly, in the UV/PMS process, the optimal PMS dose for the degradation of intI1 genes was 20 mg·L−1, while the degradation of sul1 genes gradually increased up to a PMS dose of 30 mg·L−1 [42].
For UV-based AOPs, higher UV fluence can generally result in higher ARB inactivation and ARG reduction, and the fluence-based rate constants vary for different ARB and ARGs [60], [62], [63]. In particular, it was found that the degradation efficiency of intI1 fluctuated with increasing UV fluence in the UV/PMS process [42], possibly due to its low sensitivity to direct UV photolysis. In addition to UV fluence, ARB inactivation efficiency may also be affected by UV wavelength [49], [96]. Higher bacterial inactivation rates can be achieved by UV at 265 nm than UV at 254 nm [97]. For inactivating the bacteria Campylobacter jejuni, the combination of UV at 280 and 300 nm was demonstrated to show the best performance [98]. In studies focusing on UV-based AOPs, low-pressure UV (LPUV) lamps are most commonly used, primarily because approximately 82% of LPUV radiation is released at 254 nm (near the maximum UV absorption of nucleic acids at 260 nm) [99]. Medium-pressure UV (MPUV) lamps utilize mercury vapor with a UV radiation spectrum from 200 to 300 nm. An MPUV lamp requires higher electrical input (50-250 W·cm−1) and has a shorter lifetime (4000-8000 h) than an LPUV lamp (0.5-10 W·cm−1 and 8000-10 000 h, respectively), while MPUV lamps are characterized by compact sizes and a wide range of applications [100]. Ultraviolet (UV) light-emitting diode (LED) is an emerging UV source. Although UV-LED has a relatively low wall-plug efficiency (typically less than 10%), its lack of mercury, variable wavelengths, and long lifetime make it an attractive alternative to conventional UV lamps [40], [55], [101]. Further studies on ARB and ARG control by UV-based AOPs using different UV sources are warranted.
3.3. pH
The pH of aquatic environments may vary greatly. While previous studies have described the insignificant effect of pH on ARB inactivation and ARG degradation by UV direct photolysis [41], [102], pH plays a crucial role in the formation and transformation of reactive species and thus significantly affects the efficiency of UV-based AOPs. In general, acidic conditions are favorable for the control of ARB and ARGs by UV-based AOPs.
The UV/chlorine process usually performs better in acidic environments [44], [103], mainly because of the species distribution of HOCl and OCl− (pKa = 7.5). Under acidic conditions, the dominant species of chlorine is HOCl, which has a higher quantum yield for radical generation and a lower radical scavenging capacity than OCl− [104]. Additionally, HOCl is more effective than OCl− in degrading ARGs due to its stronger oxidizing capacity [101]. When sample pH was increased from 5 to 9, a decreasing trend was observed for the reduction of sul1 and intI1 in the UV/chlorine treatment [44]. According to Eqs. (3)-(6), the scavenging of ∙OH and Cl∙ by HOCl was much slower than that by OCl−, resulting in less degradation of sul1 and intI1 under alkaline conditions. For ARGs sensitive to ClO∙, their degradation may increase with increasing pH. Yao et al. [105] reported a slight increase in the log reduction of blaNDM-1 and sul2 as pH increased from 6 to 8, mainly due to the elevated level of ClO∙. The efficiency of the UV/H2O2 process for treating ARB and ARGs generally decreases with increasing pH value under water/wastewater treatment conditions [64], [106]. For instance, the degradation of target ARGs (i.e., sul1, intI1, tetX, and tetG) during UV/H2O2 treatment of wastewater decreased as pH increased from 3.5 to 9.0 [64]. Under alkaline conditions, the formation of HO2− from H2O2/HO2− equilibrium (pKa = 11.7) becomes significant [107]. The formed HO2− can further react with H2O2 (Eq. (16)), causing a decrease in the concentration of ∙OH. Moreover, the reaction rate constant of ∙OH with HO2− is over 100 times higher than that with H2O2 (Eqs. (8), (17)), indicating the enhanced scavenging of ∙OH with increasing pH.
In the UV/PDS or UV/PMS process, ARB inactivation has been reported to decrease with increasing pH in water or wastewater treatment [84], [101], which is attributed to the enhanced decomposition of PDS or PMS and the transformation of radical species [108]. Similarly, in the pH range of 5-9, a higher reduction of target ARGs (sul1 and intI1) was achieved by the UV/PMS process under acidic conditions [42]. Under alkaline conditions, SO4∙− formed from UV-activated PDS or PMS can react with OH− to generate ∙OH (Eq. (18)). The coexistence of SO4∙− and ∙OH may result in the rapid transformation of radicals to PMS, as shown in Eq. (15) [109], which negatively affects the performance of the UV/persulfate process. Meanwhile, in UV/PDS treatment of target ARGs (i.e., blaTEM, qnrS, tetW, and tetE) at pH values of 3.4, 7.3, and 11.1, less than 20% variation in ARG degradation was observed [110], which might be due to the comparable contributions of SO4∙− and ∙OH to the ARG degradation.
pH also affects the distributions of H2CO3/HCO3− (pKa = 6.4) and HCO3−/CO32− (pKa = 10.3). CO3∙− could be produced by the reaction of ∙OH, RCS, or SO4∙− with HCO3− or CO32− (Eqs. (19), (20), (21)) [103]. The concentration of CO3∙− has been found to be much higher in the UV/chlorine process than in the UV/H2O2 process [104], which in turn is likely to be higher than that in the UV/persulfate process, due to the descending reaction rate constants of HCO3− with Cl∙, ∙OH, and SO4∙−. With increasing pH, the proportion of HCO3− and CO32− may also increase, which is unfavorable for ARB inactivation and ARG degradation because of the scavenging effect.
3.4. Water matrix
An increasing number of studies have focused on ARB inactivation and ARG degradation in real water/wastewater, where the water matrix (e.g., suspended solids, dissolved organic matter (DOM), inorganic anions, and metals) may significantly affect the efficiency of UV-based AOPs [71], [111], [112], [113]. The negative effect of suspended solids on the efficiency of UV-based AOPs due to the shielding of ARB/ARGs from UV irradiation and chemical oxidation is well acknowledged. For example, the inactivation rate of carbapenem resistant K. pneumoniae by UV/PDS was three times higher in PBS than in secondary wastewater effluent [45]. In contrast, Yoon et al. [41] found that only slightly higher UV fluence was required for ARG degradation in filtered wastewater effluent (i.e., without suspended solids) compared with that in PBS. While DOM in the water matrix is a major consumer of free radicals, it shows limited consumption of chemical oxidants and UV irradiation [49], [103]. It has been reported that ARG degradation by UV/H2O2 and UV/PDS processes was inhibited in the presence of DOM compared with the conditions without DOM [63], [95], which was due to radical scavenging by DOM. A significant decrease in radicals was observed in the UV/chlorine process in the presence of 5 mg·L−1 DOM, with reductions in the concentrations of Cl∙, ∙OH, and ClO∙ by 18%, 27%, and 99%, respectively [103]. Interestingly, when only UV irradiation was used, the presence of DOM could even enhance the degradation of ARGs, which was ascribed to the generation of reactive species via the UV photolysis of DOM [63], [95].
The presence of inorganic anions (e.g., Cl−, Br−, and SO42−) in the water matrix may affect the performance of UV-based AOPs by radical scavenging [104]. For ARB/ARGs that are not susceptible to radicals, inorganic anions have shown insignificant inhibition in their removal [72], [105]. The presence of chloride in a PMS treatment system was found to enhance the inactivation of E. coli and Bacillus spores due to the formation of HOCl/OCl− from the reaction of PMS with Cl− [114]. Studies [111], [115] have reported positive correlations of metals (e.g., Ag(I), Cu(II), Hg(II), Zn(II), and Cr(VI)) with ARG levels in aquatic environments due to the co-selection of ARGs by metals and the promotion of horizontal gene transfer at subinhibitory concentrations of metals. For instance, Cu(II) at environmentally relevant concentrations (1-100 μmol·L−1) could increase conjugation frequencies of plasmid-encoded ARGs by increasing cell membrane permeability and altering conjugative regulators [112]. In UV-based AOPs, the presence of metals with multiple redox states (e.g., Cu(I)/Cu(II), Fe(II)/Fe(III), and Mn(II)/Mn(IV)) could catalyze the decomposition of H2O2 and persulfate and enable the production of ∙OH and SO4∙−, respectively, by Fenton and Fenton-like reactions [116], [117]. The reactions of copper with H2O2, PDS, or PMS in the UV systems are described in Eqs. (22)-(25). In the Fe(II)-containing system, Fe(IV) could form and also contribute to lipid peroxidation and DNA damage [113].
4. Analysis of cost-effectiveness of the UV-based AOPs
One of the most critical factors that impact the application of a new technology is cost, which includes capital and operating costs. The capital costs of UV-based AOPs have been reported to be lower than those of other advanced technologies, such as ozonation and membrane technologies (e.g., reverse osmosis) [118]. To further evaluate the cost-effectiveness of the UV-based AOPs discussed in this study, their operating costs (including energy consumption and chemical oxidant costs) for ARB/ARG treatment were evaluated. The operating costs could be roughly estimated by the EE/O (i.e., the electrical energy required to reduce ARB/ARG levels in one cubic meter of a water sample by one order of magnitude (kW·h·m−3·order−1)) [109], [119], consisting of electrical energy for the UV source (EE/OUV) and equivalent electrical energy for the consumption of oxidants (EE/Ooxidant). The EE/O values can be calculated using Eqs. (26)-(28):
where A is the irradiated surface area (cm2), I is the UV fluence rate (mW·cm−2), t is the reaction time (h), V is the volume of the treated water samples (L), and E is the wall-plug efficiency. Since LPUV is the most commonly used UV source in studies on ARB and ARG treatment with UV-based AOPs, an E value of 0.32 was adopted according to Wan et al. [120]. C0 and Ct are the initial and final concentrations of the selected ARGs, respectively. Eqoxidant represents the electrical energy consumption required to generate one mole of an oxidant (kW·h·mol−1), and Doxidant represents the dose of the oxidant (mol·L−1). Using the average electricity cost (approximately 0.193 USD·(kW·h)−1 [121] and chemical prices from manufacturers and suppliers on a well-known e-commerce platform†(† https://www.alibaba.com/.), we calculated the Eqoxidant values for chlorine, H2O2, PDS, and PMS to be 4.64, 3.44, 4.98, and 7.79 kW·h·mol−1, respectively. Using the UV fluences and chemical doses reported in previous studies for ARB/ARG treatment (Table 2, Table 3, Table 4), we determined the costs for energy consumption and chemical oxidants in different UV-based AOPs.
Fig. 3 shows the EE/OUV and EE/Ooxidant of different UV-based AOPs for ARB/ARG treatment in PBS or wastewater. A higher EE/O is indicative of a low energy efficiency of the corresponding treatment [119]. The median EE/O values of UV/chlorine, UV/H2O2, and UV/persulfate processes for ARG degradation in PBS and wastewater were averagely 7.7 and 29.7 times larger than those for ARB inactivation, respectively. This indicates that, compared with ARG degradation, UV-based AOPs are generally very effective in ARB inactivation. In the PBS solution, the UV/persulfate process was the most cost-effective UV-based AOP (Fig. 3(a)). Both the median values of EE/OUV and EE/Ooxidant in the UV/persulfate process for ARG degradation (0.064 and 0.324 kW·h·m−3·order−1, respectively) were lower than the corresponding values in UV/chlorine (0.334 and 0.408 kW·h·m−3·order−1, respectively) and UV/H2O2 (0.266 and 1.720 kW·h·m−3·order−1, respectively) processes. In the PBS solution without interferences from the water matrix, the relatively high EE/Ooxidant of the UV/H2O2 process might be attributed to the lower absorbance coefficients of H2O2 (18.6 mol−1·cm−1) than those of other oxidants (e.g., 59 mol−1·cm−1 for HOCl and 27.5 mol−1·cm−1 for PDS) at UV 254 nm [55], [122]. The UV/H2O2 process usually requires excess H2O2 because only 5%-10% of the dosed H2O2 is consumed by UV photolysis, which leads to an increase in chemical costs [122]. Additional chemicals are required to remove the remaining H2O2; accordingly, the operation of the UV/H2O2 process is relatively complex and less effective in terms of energy utilization [123]. However, in a wastewater matrix, the UV/chlorine process showed higher energy efficiency than other UV-based AOPs (Fig. 3(b)). Although the median EE/OUV of the UV/persulfate process for ARG degradation (0.179 kW·h·m−3·order−1) was lower than that of the UV/chlorine and UV/H2O2 processes (0.421 and 0.277 kW·h·m−3·order−1, respectively), the median EE/Ooxidant increased significantly to 9.68 kW·h·m−3·order−1. The higher energy efficiency of UV/chlorine in wastewater might result from the higher disinfection efficacy of chlorine than H2O2 and persulfate [77], [78] and the comparatively high selectivity of RCS over ∙OH and SO4∙− [124]. Similar results have been reported for cost-effectiveness analysis related to the removal of emerging contaminants in wastewater [118], [125]. Guo et al. [119] reported that the electrical energy required for organic pollutant removal was lower in the UV/chlorine process than in the UV/H2O2 process. With regard to emerging contaminant degradation, the UV/chlorine process resulted in savings of 30%-75% in electrical energy compared with the UV/H2O2 process, resulting in a remarkable 30%-50% reduction in operating costs and a potential reduction in capital costs owing to the smaller number of reactor chambers required for the UV/chlorine process [118]. A previous review [126] summarized the energy efficiency-of UV-based AOPs in the degradation of pharmaceuticals and indicated that the EE/O values followed the order of EE/OUV/catalyst > EE/OUV/H2O2 > EE/OUV/persulfate > EE/OUV/chlorine.
It is worth noting that in the EE/O analysis, the number of available data (n) for the UV/chlorine (n = 9) and UV/persulfate (n = 13) processes was smaller than that for the UV/H2O2 process (n = 27), possibly because UV/chlorine and UV/persulfate are new AOPs. More studies on ARG degradation by different UV-based AOPs are thus warranted to further improve the EE/O analysis. Additionally, despite the cost-effectiveness of the UV/chlorine process in treating ARB and ARGs, halogenated disinfection byproducts (DBPs) can be readily formed from reactions of DOM with chlorine and RCS in this process [104], [127], [128]. UV/chlorine treatment of drinking water and wastewater has been reported to generate more halogenated DBPs than chlorination alone [128], [129]. As many halogenated DBPs are cytotoxic, genotoxic, and developmentally toxic [130], [131], [132], [133], [134], [135], increased toxicity of wastewater effluent from UV/chlorine treatment has been observed [136] as a consequence of DBP formation. Concerns about the formation of toxic halogenated byproducts should be thoroughly evaluated before the application of the UV/chlorine process.
5. Conclusions and prospects
UV-based AOPs (e.g., UV/chlorine, UV/H2O2, and UV/persulfate) are considered promising control technologies to address the pressing health concerns associated with ARB and ARGs. Due to the generation of multiple reactive species, these integrated AOP systems show significantly higher treatment efficiencies than direct UV photolysis or chemical oxidation alone. However, the specific reactive species mainly responsible for ARB inactivation and ARG degradation depend on the target bacterial strains and gene types. The efficiency of UV-based AOPs can be influenced by factors such as types and conditions of target ARB and ARGs, chemical dose and UV fluence, pH, and water matrix properties. The combined effects of these influencing factors further complicate the control of ARB and ARGs by UV-based AOPs. To evaluate the feasibility of using UV-based AOPs for ARB and ARG control, the EE/O values (considering the electrical energy for LPUV lamps and chemical oxidants) of the UV/chlorine, UV/H2O2, and UV/persulfate processes were analyzed. According to the EE/O results, while the UV/persulfate process showed the best energy efficiency in PBS solution, the UV/chlorine process was most efficient in real wastewater. The different EE/O rank orders highlighted the influence of the water matrix on the efficiency of UV-based AOPs. Additionally, the type of UV lamps used (e.g., LPUV versus MPUV and UV-LED) and the treatment capacity (e.g., laboratory-scale versus pilot- and full-scale) may also affect the cost evaluation for different UV-based AOPs, necessitating further investigation and systematic assessment. Notably, although the UV/chlorine process was more cost-effective in controlling ARB and ARGs than other UV-based AOPs, the formation of toxic byproducts during the UV/chlorine process may counteract its benefits and should be carefully evaluated.
Acknowledgments
The work described in this paper was supported by grants from the Research Grants Council of the Hong Kong SAR, China (T21-705/20-N and 16210221).
Compliance with ethics guidelines
Jiarui Han, Wanxin Li, Yun Yang, Xuanwei Zhang, Siyu Bao, Xiangru Zhang, Tong Zhang, and Kenneth Mei Yee Leung declare that they have no conflict of interest or financial conflicts to disclose.
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