Direct and Activated Chlorine Dioxide Oxidation for Micropollutant Abatement: A Review on Kinetics, Reactive Sites, and Degradation Pathway

Abstract

Recently, ClO₂-based oxidation has attracted increasing attention to micropollutant abatement, due to high oxidation potential, low disinfection byproduct (DBPs) formation, and easy technical implementation. However, the kinetics, reactive sites, activation methods, and degradation pathways involved are not fully understood. Therefore, we reviewed current literature on ClO₂-based oxidation in micropollutant abatement. In direct ClO₂ oxidation, the reactions of micropollutants with ClO₂ followed second-order reaction kinetics (k_app = 10⁻³–10⁶ M⁻¹ s⁻¹ at neutral pH). The k_app depends significantly on the molecular structures of the micropollutant and solution pH. The reactive sites of micropollutants start with certain functional groups with the highest electron densities including piperazine, sulfonyl amido, amino, aniline, pyrazolone, phenol groups, urea group, etc. The one-electron transfer was the dominant micropollutant degradation pathway, followed by indirect oxidation by superoxide anion radical (O₂^•−) or hydroxyl radical (^•OH). In UV-activated ClO₂ oxidation, the reactions of micropollutants followed the pseudo-first-order reaction kinetics with the rates of 1.3 × 10⁻⁴–12.9 s⁻¹ at pH 7.0. Their degradation pathways include direct ClO₂ oxidation, direct UV photolysis, ozonation, ^•OH-involved reaction, and reactive chlorine species (RCS)-involved reaction. Finally, we identified the research gaps and provided recommendations for further research. Therefore, this review gives a critical evaluation of ClO₂-based oxidation in micropollutant abatement, and provides recommendations for further research.

1. Introduction

The micropollutants also known as contaminants of emerging concerns (CECs) are comprised of various anthropogenic and natural compounds, such as pharmaceuticals and personal care products (PPCPs), endocrine disruptors, and pesticides [1,2]. Their presence in natural and engineered systems, even at trace concentrations (ng L⁻¹ to μg L⁻¹), has attracted significant attention because of their toxic, persistent, bioaccumulative properties [3]. Due to increased industrialization and urbanization, many micropollutants are widely used, and eventually end up in different types of wastewaters. Unfortunately, traditional wastewater treatment plants (WWTPs) are not explicitly designed for micropollutant abatement, resulting in WWTPs being one of the significant sources of micropollutants in surface water [4]. Until now, various techniques have been proposed for micropollutant abatement in WWTPs, including activated carbon/biochar adsorption [5], advanced oxidation processes (AOPs) [6,7], and membrane filtration [8]. AOPs have attracted growing attention among these techniques because of their simple operation, high removal efficiency, and rapid oxidation.

AOPs enable an approach combining two individual processes of disinfection and decontamination with improved cost-effectiveness [9,10]. Chlorine Dioxide (ClO₂) is a green disinfectant/oxidant and was listed as an A1-level, safe, and efficient disinfectant by the World Health Organization (WHO) [11]. It has been prevalently used as a drinking water disinfectant alternative to chlorine (Cl₂) due to its effectiveness in pathogen inactivation and limited formation of halogenated disinfection byproducts (DBPs), such as trihalomethanes (THMs) and haloacetic acids (HAAs) [12]. In addition, the products of ClO₂ disinfection/oxidation consist 50–70% of chlorite (ClO₂⁻) and 30–50% of chlorate (ClO₃⁻) and chloride (Cl⁻) [13,14]. Controlling the levels of these ClO₂ residuals is critical for successfully implementing ClO₂ disinfection/oxidation.

Recently, ClO₂-based oxidation for micropollutant abatement has attracted increased attention due to its advantages of strong oxidation, low DBPs formation, and easy technical implementation. ClO₂ can effectively oxidize micropollutants with electron-rich functional groups such as aniline, phenolic, aromatic, and tertiary amine groups [15,16]. ClO₂ is typically transformed into ClO₂⁻ through a single-electron oxidation process [17]. Recent studies used external energy to activate ClO₂ to produce reactive species, resulting in improved micropollutant abatement. For example, the excellent performance of the co-exposure of ClO₂ and ultraviolet radiation (UV) was reported in micropollutant abatement due to the high yield of reactive species [18,19].

ClO₂-based oxidation includes direct and activated ClO₂ oxidations. Though studies on ClO₂-based oxidation, especially on the direct ClO₂ oxidation, have been increasing over the past decade, there is still a limited understanding of these processes on micropollutant abatement, such as their kinetics, reactive sites, activation methods, and degradation pathways. The existing review about ClO₂ primarily focused on the reaction with (in)organic compounds in water treatment [17], pathogenic microbe inactivation in water treatment [20], antimicrobial food packaging [21], disposable ClO₂ wipes [22], and postharvest handling and food storage [23]. To the best of our knowledge, there is no comprehensive review on ClO₂-based oxidation in micropollutant abatement. Therefore, providing a comprehensive review of this technology is crucial for future research and application. In this review, we emphatically discussed (1) ClO₂ properties; (2) reaction kinetics, reactive sites, and degradation pathways in the directed ClO₂ oxidation; and (3) reaction kinetics and degradation pathways in the UV-activated ClO₂ oxidation.

2. ClO₂ Physicochemical Properties

ClO₂ is a green-yellowish gas and has a pungent odor similar to Cl₂. It is one of the few compounds in nature that exist almost entirely as monomeric free radicals due to a single unpaired electron [24]. The molecular weight and the standard oxidation state of Cl atoms in ClO₂ are 67.46 and +4, respectively. ClO₂ has a boiling point of 11 °C, a melting point of −59 °C, a density of 1.64 g mL⁻¹ (liquid) at 0 °C [25], a water solubility of 3.0 g L⁻¹ at 25 °C [17], and pKa value of 3.0. ClO₂ is strongly soluble in water and does not hydrolyze to any appreciable extent but remains in solution as a dissolved gas [26]. ClO₂ in aqueous solutions is quite stable when protected from light and kept cool, well-sealed, and slightly acidified (pH = 6). The ultraviolet absorption spectrum of ClO₂ solutions has broadband with a peak at 359 nm and a molar extinction coefficient of ~1250 M⁻¹ cm⁻¹ [27].

ClO₂ has a relatively short half-life and is highly volatile and explosive under concentrations of >10% in the air [28]. ClO₂ solution under concentrations of <~10 g L⁻¹ will not produce sufficiently high vapor pressure for an explosive hazard. In water treatment practice, the concentrations of concentrated ClO₂ solution rarely exceed 4 g L⁻¹. Furthermore, ClO₂ cannot be compressed, stored, or transported under pressure and must be generated on-site [29]. Compared with the electrolysis method, the chemical method is more mature for ClO₂ production, which refers to the reactions of sodium chlorite (NaClO₂) or sodium chlorate (NaClO₃) with Cl₂, hydrochloric acid (HCl), or peroxydisulfate (H₂S₂O₈) (Equations (1)–(3)). The reaction of NaClO₂ with an acid, such as HCl, has become an increasingly common method for ClO₂ production due to the operational difficulty and safety concerns of handling Cl₂ gas. Noted, to produce the same mass weight of ClO₂, hydrochloric-based ClO₂ production (Equation (2)) uses 1.25 times more NaClO₂ than chlorine-based (Equation (1)) or peroxydisulfate-based (Equation (3)) ClO₂ production.

2NaClO₂ + Cl₂ → 2ClO₂ + 2NaCl

(1)

5NaClO₂ + 4HCl → 4ClO₂ + 5NaCl + 2H₂O

(2)

2NaClO₂ + 4Na₂S₂O₈ → 2ClO₂ + 2Na₂SO₄

(3)

In the water, ClO₂ reacts first with other compounds to form ClO₂⁻ through a one-electron transfer reaction (Equation (4)), with the redox potential of 0.936 V [30]. The second reaction of the formed ClO₂⁻ transforming to Cl⁻ by gaining four electrons does not occur readily, due to the low reactivity of ClO₂⁻ (Equation (5)). In practice, fast oxidation predominates, and therefore, ClO₂⁻ will be the significant byproduct during ClO₂ disinfection/oxidation [31]. ClO³⁻ will be another byproduct because of its presence in proprietary solutions of ClO₂. ClO₂ accepts five electrons when thoroughly reduced to Cl⁻, while Cl₂ accepts two electrons from the oxidation compounds (Equations (6) and (7)). Therefore, the oxidative capacity of ClO₂ is also approximately 2.5 times of Cl₂ on a weight basis.

ClO₂ + e⁻ → ClO₂⁻

(4)

ClO₂⁻ + 2H₂O + 4e⁻ → Cl⁻ + 4OH

(5)

ClO₂ + 2H₂O + 5e⁻ → Cl⁻ + 4OH⁻

(6)

Cl₂ + 2e⁻ → 2Cl⁻

(7)

3. Direct ClO₂ Oxidation

3.1. Reaction Kinetics

The reaction kinetics between ClO₂ and micropollutants can be well described by second-order kinetic models (Equations (8) and (9)), referring to a first-order model in ClO₂ concentration and a first-order model in micropollutant concentration [32,33].

ClO₂ + MP → product

(8)

$${\displaystyle \frac{\mathrm{d}\left[\mathrm{MP}\right]\mathrm{tot}}{\mathrm{dt}}}=-{\mathrm{k}}_{\mathrm{app}}[\mathrm{MP}{]}_{\mathrm{tot}}[{\mathrm{ClO}}_2]$$

(9)

where MP is an organic micropollutant; k_app is the apparent second-order rate constant for the overall reaction; [ClO₂] and [MP]_tot is the ClO₂ and MP concentration, respectively.

The reaction kinetics of antibiotics with ClO₂ depends on their molecular structures and pH. The k_app of antibiotics ranged from 1.2 to 1.3 × 10⁶ M⁻¹ s⁻¹ at neutral pH, with a general order of tetracyclines (10⁵–10⁶) > triclosan (10⁴–10⁵) > sulfonamides (10³–10⁴) > macrolides (10¹–10²) > fluoroquinolones (1–10¹) (

Table 1. Second-order rate constants (k_app) in the reaction of ClO₂ with micropollutants.

Compounds	kapp (M−1 s−1)	pH	T (°C)	References
fluoroquinolones
Ciprofloxacin	1.2	7.0	-	[37]
Ciprofloxacin	7.9	7.0	22	[34]
Norfloxacin	1.3 × 101	7.0	22	[34]
Lomefloxacin	6.8	7.0	22	[34]
Ofloxacin	7.8 × 101	7.0	22	[34]
Pipemidic acid	1.5	7.0	22	[34]
Enrofloxacin	6.3 × 101	7.0	22	[34]
tetracyclines
Tetracycline	1.3 × 106	7.0	22	[35]
Oxytetracycline	1.2 × 106	7.0	22	[35]
Chlorotetracycline	3.2 × 105	7.0	22	[35]
Iso-chlorotetracycline	2.2 × 105	7.0	22	[35]
sulfonamides
Sulfamethoxazole	6.7 × 103	7.0	20	[15]
Sulfamethoxazole	7.9 × 103	7.0	-	[37]
Sulfamethoxazole	6.1 × 103	7.0	20	[36]
Sulfamethizole	3.9 × 103	7.0	20	[36]
Sulfadimethoxine	4.4 × 103	7.0	20	[36]
Sulfamethazine	4.1 × 103	7.0	20	[36]
Sulfamerazine	5.6 × 103	7.0	20	[36]
Sulfathiazole	2.6 × 104	7.0	20	[36]
macrolides
Roxithromycin	2.2 × 102	7.0	20	[15]
Roxithromycin	8.8 × 101	7.0	-	[37]
triclosan
Triclosan	7.1 × 104	~7.0	rt	[38]
Triclosan	6.3 × 105	7.0	-	[37]
antipyretic analgesics
Antipyrine	4.8 × 10−1	7.0	25	[39]
Propylphenazone	>100	7.4	20	[15]
Propylphenazone	1.1 × 101	7.0	25	[40]
Naproxen	6.1 × 102	7.0	-	[37]
Naproxen	10–100	7.4	20	[15]
Aminopyrine	1.3 × 105	7.0	25	[40]
Aminopyrine	>100	7.4	20	[15]
Diclofenac	1.1 × 104	7.0	20	[15]
Diclofenac	1.5 × 103	7.0	25	[33]
Diclofenac	1.1 × 104	7.0	-	[37]
Acetaminophen	2.1 × 105	7.0	-	[37]
Fenoprofen	<1	7.4	20	[15]
Ibuprofen	<0.1	8.0	-	[41]
β-blockers
Atenolol	~1	8.0	-	[41]
Metoprolol	1.3	8.0	20	[42]
antiepileptics
Carbamazepine	<0.1	8.0	-	[41]
psychostimulants
Caffeine	<1	7.4	20	[15]
antineoplastics
Ifosfamide	<1	7.4	20	[15]
Cyclophosphamide	<1	7.4	20	[15]
lipid regulators
Gemfibrozil	5.9 × 101	7.0	-	[37]
Gemfibrozil	<10	7.4	20	[15]

-: not available; rt: room temperature.

). There is the aniline moiety in tetracyclines and sulfonamides and phenol moiety in triclosan, but the alkyl amine moiety in macrolides and fluoroquinolones. The results suggested that antibiotics with the aniline and phenol groups may be more vulnerable to ClO₂ attack than those with the alkyl amine. Furthermore, enrofloxacin and ofloxacin with tertiary amines on piperazine moieties reacted faster with ClO₂ than other fluoroquinolones with secondary amines on piperazine moieties [34]. However, the reactivity of the dimethylamino group in tetracycline to ClO₂ is higher than that of trimethylamine but lower than that of N,N-dimethylaniline [35]. The k_app of antibiotics was related to pH as well. A kinetic study demonstrated that the k_app of ciprofloxacin (belonging to fluoroquinolones) increased by more than three orders of magnitude from pH 4.48 to 9.55 [34]. Similarly, the k_app increased by more than 4 to 6 and 1.6 to 2.2 orders of magnitude from pH 2.5 to 10.5 and from 4.0 to 9.5 for the ClO₂ oxidation of tetracyclines and sulfonamides, respectively [35,36]. The large variation with pH could be attributed to the varying reactivity of antibiotic acid-base species toward ClO₂. An increase in pH led to a larger fraction of the deprotonated species (A⁻), thus facilitating the reaction of antibiotics with ClO₂. Similar trends were also observed in fluoroquinolones [34] and tetracyclines [35], indicating that the deprotonation of these antibiotics as pH increases considerably favors their oxidation by ClO₂.

In addition, the direct ClO₂ oxidation was applied for degrading other PPCPs such as antipyretic analgesics, β-blockers, antiepileptics, psychostimulants, antineoplastics, and lipid regulator. The k_app of antipyretic analgesics ranged from 4.8 × 10⁻¹ to 2.1 × 10⁵ M⁻¹ s⁻¹ at neutral pH (Table 1). Among them, aminopyrine, diclofenac, and acetaminophen had the highest k_app due to the tertiary amine, aniline, and phenol moiety in their molecule, respectively. The second highest k_app was observed in propylphenazone with the heterocyclic amine moiety and naproxen with substituted benzene moiety. The remaining studied PPCPs were less reactivity towards ClO₂ (Table 1). These results implied that ClO₂ is a highly selective oxidant with respect to micropollutants with specific functional groups such as aniline, phenolic moieties, the second and tertiary amine, heterocyclic amine, aromatic nucleus.

3.2. Reactive Sites

Reactive sites of micropollutants during ClO₂ oxidation are determined by functional groups with the highest electron densities due to the one-electron transfer mechanism. As for PPCPs, the main reactive sites include piperazine group, sulfonyl amido group, amino group, aniline group, pyrazolone group and phenol group (Table 2). The N4 atom in the piperazine ring of fluoroquinolones was the specific site to be attacked by ClO₂ [34]. Similarly, He et al. [43,44] found the tertiary N4 amines and the secondary N4 amines with the highest 2FED_HOMO² value in the piperazinyl group as the most vulnerable sites in the reactions between ClO₂ and the fluoroquinolones of fleroxacin and enoxacin. The sulfonyl amido-nitrogen of sulfonamides could be the main reaction site toward ClO₂ [36]. The reaction of three representative β-lactam antibiotics with ClO₂ starts with a single-electron transfer from the lone electron pair of the amino group to ClO₂ [45]. ClO₂ reacts with tetracyclines predominantly in the unprotonated dimethylamino group and deprotonated phenolic-diketone group [35]. Furthermore, the reactive site of triclosan was the phenol group during ClO₂ oxidation [37]. For antipyretic analgesics, the aniline group in diclofenac was the reactive site and acted as the electron-rich moieties during ClO₂ oxidation [37]. The N2 atom on the pyrazolone ring of antipyrine was vulnerable under the electrophilic reaction of ClO₂ due to its high electron cloud density [39]. However, the C=C double bond on the pyrazolone ring of isopropylphenazone and aminopyrine were the most reactive sites toward ClO₂ [40].

The main reactive sites of pesticides are the urea group and aromatic benzene ring of phenylurea and sulfonylurea herbicides, the sulfur center of ametryn and methiocarb, the amide group and the phosphinothioyl group of organophosphorus pesticides. For example, the primary attack on two phenylurea herbicides of diuron and chlortoluron by ClO₂ might be the electron-rich nitrogen atom on the ureic side-chain [46]. However, the aromatic benzene ring of isoproturon is vulnerable to the attack of ClO₂ [47]. Additionally, the degradation of two sulfonylurea herbicides of nicosulfuron and thifensulfuron methyl started with an attack on the urea groups by ClO₂ [48]. The main reactive site of ametryn herbicide and methiocarb pesticide during ClO₂ oxidation was the sulfur center in their molecules [47,49]. ClO₂ oxidation of two organophosphorus pesticides started with an attack on the amide group of azamethiphos and the phosphinothioyl group of dimethoate [16].

Table 2. Reactive sites and degradation pathways based on the intermediate products in the reaction of ClO₂ with micropollutants.

Micropollutants	Molecular Structure	Reactive Sites	Pathways	References
β-lactam antibiotics
Amoxicillin		amino group	pathway: β-lactam ring breaking	[45]
Cefadroxil
fluoroquinolones
Ciprofloxacin		piperazine’s N4 atom	pathway: dealkylation, hydroxylation and intramolecular ring closure at the piperazine moiety, and the quinolone ring mostly intact	[34]
Norfloxacin
Enoxacin			pathway: piperazine group cleavage, the decarboxylating quinolone ring, and hydroxylation	[44]
Fleroxacin			pathway: the cleavage of the piperazine ring and the decarboxylation and chlorination of the quinolone ring	[32]
sulfonamides
Sulfamethoxazole		sulfonyl amido-nitrogen	pathway: breakage of S-N and C-S bonds and hydroxylation of aniline group	[36]
tetracyclines
Tetracycline		dimethylamino groups	pathway: (hydr)oxylation and breakage of tetracycline molecules	[35]
triclosan
Triclosan		phenol group a	pathway: the closure of the phenolic ring, chlorination of the phenolic ring and cleavage of the ether bond	[38]
antipyretic analgesics
Diclofenac		aniline group a	pathway: decarboxylation, hydroxylation and chlorination of the phenylacetic acid moiety, and further C-N bond cleavage	[50]
Antipyrine		pyrazolone’s N2 atom	pathway: chlorination substitution, ring-opening reaction and de-carbonyl reaction of the pyrazolone ring	[39]
Iso-propylphenazone		pyrazolone’s C=C	pathway: C=C cleavage, ring opening reaction and de-carbonyl reaction of the pyrazolone ring	[40]
Aminopyrine
antidepressant
Venlafaxine		-	pathway: dehydration, demethylation and cleavage of the molecular structure	[51]
phenylurea herbicides
Fenuron		urea group	pathway: electrophilic substitution and cleavage of the urea group products: a chloro-quinone product and an urea derivative	[52]
Isoproturon		aromatic benzene ring	pathway: aromatic-ring hydroxylated substituted derivatives	[47]
Chlortoluron		nitrogen atom on the ureic side-chain	pathway: radical intermediates formation, hydroxylation reactions and cleavage of the N–C bond on the ureic side-chain	[46]
Diuron			pathway: hydroxylation reactions and cleavage of the N–C bond on the ureic side-chain, dechloridation of the benzene ring
sulfonylurea herbicides
Nicosulfuron		urea group	pathway: the urea group breaking	[48]
herbicide
Ametryn		sulfur center	pathway: oxidation reactiveproduct: the sulphoxide derivative	[47]
carbamate pesticides
Methiocarb		sulfur center	pathway: oxidation reaction products: methiocarb sulfoxidemethiocarb sulfone	[49]
organophosphorus pesticides
Azamethiphos		amide group	pathway: the breaking of amide group or S–C bond	[16]
Dimethoate		the phosphinothioyl group	pathway: the breaking of S–C bond, oxidation of the phosphinothioyl group

^a [37]; -: not available.

Table 2. Reactive sites and degradation pathways based on the intermediate products in the reaction of ClO₂ with micropollutants.

Micropollutants	Molecular Structure	Reactive Sites	Pathways	References
β-lactam antibiotics
Amoxicillin		amino group	pathway: β-lactam ring breaking	[45]
Cefadroxil
fluoroquinolones
Ciprofloxacin		piperazine’s N4 atom	pathway: dealkylation, hydroxylation and intramolecular ring closure at the piperazine moiety, and the quinolone ring mostly intact	[34]
Norfloxacin
Enoxacin			pathway: piperazine group cleavage, the decarboxylating quinolone ring, and hydroxylation	[44]
Fleroxacin			pathway: the cleavage of the piperazine ring and the decarboxylation and chlorination of the quinolone ring	[32]
sulfonamides
Sulfamethoxazole		sulfonyl amido-nitrogen	pathway: breakage of S-N and C-S bonds and hydroxylation of aniline group	[36]
tetracyclines
Tetracycline		dimethylamino groups	pathway: (hydr)oxylation and breakage of tetracycline molecules	[35]
triclosan
Triclosan		phenol group a	pathway: the closure of the phenolic ring, chlorination of the phenolic ring and cleavage of the ether bond	[38]
antipyretic analgesics
Diclofenac		aniline group a	pathway: decarboxylation, hydroxylation and chlorination of the phenylacetic acid moiety, and further C-N bond cleavage	[50]
Antipyrine		pyrazolone’s N2 atom	pathway: chlorination substitution, ring-opening reaction and de-carbonyl reaction of the pyrazolone ring	[39]
Iso-propylphenazone		pyrazolone’s C=C	pathway: C=C cleavage, ring opening reaction and de-carbonyl reaction of the pyrazolone ring	[40]
Aminopyrine
antidepressant
Venlafaxine		-	pathway: dehydration, demethylation and cleavage of the molecular structure	[51]
phenylurea herbicides
Fenuron		urea group	pathway: electrophilic substitution and cleavage of the urea group products: a chloro-quinone product and an urea derivative	[52]
Isoproturon		aromatic benzene ring	pathway: aromatic-ring hydroxylated substituted derivatives	[47]
Chlortoluron		nitrogen atom on the ureic side-chain	pathway: radical intermediates formation, hydroxylation reactions and cleavage of the N–C bond on the ureic side-chain	[46]
Diuron			pathway: hydroxylation reactions and cleavage of the N–C bond on the ureic side-chain, dechloridation of the benzene ring
sulfonylurea herbicides
Nicosulfuron		urea group	pathway: the urea group breaking	[48]
herbicide
Ametryn		sulfur center	pathway: oxidation reactiveproduct: the sulphoxide derivative	[47]
carbamate pesticides
Methiocarb		sulfur center	pathway: oxidation reaction products: methiocarb sulfoxidemethiocarb sulfone	[49]
organophosphorus pesticides
Azamethiphos		amide group	pathway: the breaking of amide group or S–C bond	[16]
Dimethoate		the phosphinothioyl group	pathway: the breaking of S–C bond, oxidation of the phosphinothioyl group

^a [37]; -: not available.

3.3. Degradation Pathways

One-electron transfer was the dominant degradation pathway in the direct ClO₂ oxidation of micropollutants, followed by indirect oxidation by oxygen species such as superoxide anion radical (O₂^•−) or hydroxyl radical (^•OH) (Figure 1). In detail, the one-electron transfer oxidation pathway refers to: (1) ClO₂ attacks the atom of the micropollutant with the highest electron density or the strongest electron-donating and takes away an electron from the atom to make micropollutant forming a radical intermediate (MP^•), and is thus reduced to ClO₂⁻ (Equation (10); it is the rate-determining step) and (2) MP^• combines with another ClO₂ to form degradation products by undergoing molecular rearrangement and binding to itself or ClO₂ (Equation (11)). During the direct ClO₂ oxidization of MP^•, the oxidant is initially reduced to ClO, then to HClO, and subsequently to Cl⁻. Therefore, four tentative routes occur through the oxygen transfer process and potentially contribute to direct ClO₂ oxidization of micropollutants (Equations (12)–(15)), which were confirmed by the identification of decarbonyl-MP, hydroxyl-MP, chloro-MP, and even ring rupture of MP [38,53,54].

MP + ClO₂ → MP^• + ClO₂⁻

(10)

MP^• + ClO₂ → products

(11)

MP^• + ClO₂ → MP − OH + ClO

(12)

MP^• + ClO₂ → decarbonyl − MP + ClO

(13)

MP^• + H⁺ + ClO → MP⁺ + HOCl

(14)

MP^• + HOCl → MP − Cl + H₂O

(15)

ClO₂ oxidation of PPCPs mainly led to the ring-opening reaction, dealkylation, decarboxylation, hydroxylation, and chlorination. The cleavage of the β-lactam ring in the molecules of penicillin, amoxicillin, and cefadroxil was observed in the ClO₂ oxidation of β-lactam antibiotics [45]. ClO₂ oxidation of fluoroquinolones, ciprofloxacin, and norfloxacin, led to dealkylation, hydroxylation, and intramolecular ring closure at the piperazine moiety but left the quinolone ring mostly intact [34]. Similarly, the primary and initial step in the ClO₂ oxidation of enoxacin and fleroxacin was the cleavage of the piperazine ring [32,44]. The decarboxylation and hydroxylation or chlorination of the quinolone ring occurred in enoxacin and fleroxacin, whereas the quinolone ring was unreactive of ciprofloxacin and norfloxacin. The cleavage of S−N and C−S bonds and the hydroxylation of aniline moiety were the major degradation pathways of sulfamethoxazole [36]. Furthermore, (hydr)oxylation and breakage of tetracycline were observed during the ClO₂ oxidation [35]. The pyrazolone ring-opening reaction caused by C=C double bond cleavage and further de-carbonyl reaction were the main degradation pathways of three antipyretic analgesics of antipyrine, isopropylphenazone, and aminopyrine [39,40]. ClO₂ oxidation of triclosan involved the cleavage of the ether link, chlorination of the phenolic ring, and ring closure [38]. The transformation pathways of venlafaxine included dehydration, demethylation, and cleavage of the molecule during ClO₂ oxidation [51].

ClO₂ reacts with commonly used phenylurea herbicides and sulfonylurea herbicides predominantly by the cleavage of the urea group and hydroxylated substitutes of the aromatic ring. ClO₂ oxidation of phenylurea herbicides of chlortoluron and diuron was subjected to steps including radical intermediates formation, hydroxylation reactions, and cleavage of the N−C bond on the ureic side-chain [46]. Isoproturon, a phenylurea-derivative, reacts with ClO₂ to form aromatic-ring hydroxylated substituted derivatives [47]. The urea group in sulfonylurea herbicide of nicosulfuron reacts firstly with ClO₂, resulting in breaking one bond and forming two degradation products of 2-(Nformylsulfamoyl)-N,N-dimethylnicotinamide and 4,6-dimethoxypyrimidin-2-amine [48].

Additionally, ClO₂ oxidation of other pesticides with the sulfur or phosphinothioyl center in their molecules led to sulfoxide and sulfone, ring rupture, hydroxylation, and decarbonyl products. Ametryn (R-S-CH3) reacted with ClO₂ forming the sulfoxide derivative (R-SO-CH3) [47]. Similarly, during the ClO₂ oxidation of the carbamate pesticide of methiocarb (MC), methiocarb sulfoxide and methiocarb sulfone were generated by losing HClO₂ and HOCl from the intermediate adduct of MC-ClO₂-OH, respectively, which was first formed around the sulfur center of methiocarb [49]. Pergal, et al. [16] studied the ClO₂ oxidation of two organophosphorus pesticides and found the successive attack on the amide group and the sulfide group in azamethiphos, leading to the break of the amide group and S–C bond, and the hydroxylation of the phosphinothioyl and then decarbonyl in dimethoate.

The indirect oxidation pathway of ClO₂ with micropollutants includes (1) the formation of O₂^•− by concurrently transferring an electron from MP^• to dissolved oxygen in solution (Equation (16)), (2) the reaction of O₂^•− with water with the formation of ^•OH (Equation (17)), and (3) the degradation of micropollutant by the formed O₂^•−/^•OH (Equation (18)). For example, two major degradation pathways of diclofenac (DCF) were proposed as (1) direct ClO₂ oxidation through one-electron transfer and (2) indirect O₂^•− oxidation by concurrently transferring an electron from DCF^• to dissolved oxygen [33].

MP^• + O₂ → MP⁺ + O₂^•−

(16)

O₂^•− + H₂O → ^•OH

(17)

MP + O₂^•−/^•OH → products

(18)

4. UV-Activated ClO₂ Oxidation

4.1. Reaction Kinetics

The micropollutant abatements were generally enhanced by combining ClO₂ with shortwave ultraviolet light (UVC), which follows the pseudo-first-order reaction kinetics with the pseudo-first-order rates (k_obs) of 1.3 × 10⁻⁴–9.8 × 10⁻³ s⁻¹ at pH 7.0 (

Table 3. Summary of research studying the removal of micropollutants by the UV-activated ClO₂ oxidation.

Micropollutants	C0	ClO2	UV Light	Light Intensity	Reaction pH	kobs (s−1)	Removal Rate (%)	References
Triclosan	0.3 mg L−1	0.5 mg L−1	UVC	6.5 μW cm−2	~7.0	-	>99	[55]
Trimethoprim	1 μg L−1	1.4 mg L−1	UVC	207 mJ cm−2	7.0	5.7 × 10−4	14.4–100.0	[18]
Iopromide	1 μg L−1	1.4 mg L−1	UVC	207 mJ cm−2	7.0	1.2 × 10−3	14.4–100.0	[18]
Caffeine	1 μg L−1	1.4 mg L−1	UVC	207 mJ cm−2	7.0	1.3 × 10−4	14.4–100.0	[18]
Ciprofloxacin	1 μg L−1	1.4 mg L−1	UVC	207 mJ cm−2	7.0	9.8 × 10−3	14.4–100.0	[18]
Iopamidol	10 μM	200 μM	UVC	2.4 mW cm−2	7.0	4.4 × 10−3	74.9	[57]
Micropollutants a	1 μM	5 mg L−1	UVA	0.3 mW cm−2	7.0	3.8 × 10−4 to 12.9	7–100	[19]

-: not available; ^a 19 micropollutants: iopromide, trimethoprim, caffeine, 17α-ethynylestradiol, 17β-estradiol, estrone, diclofenac, sulfamethoxazole, gemfibrozil, naproxen, ofloxacin, roxithromycin, carbamazepine, metoprolol, atenolol, metronidazole, bezafibrate, clofibric acid, ibuprofenreported.

). For example, more than 99% of triclosan was degraded under co-exposure to UVC irradiation and ClO₂ [55]. Four micropollutants (i.e., trimethoprim, iopromide, caffeine, and ciprofloxacin) were degraded by 14.4–100.0% in UVC/ClO₂, with the corresponding k_obs following an order of ciprofloxacin (9.8 × 10⁻³ s⁻¹) > iopromide (1.2 × 10⁻³ s⁻¹) > trimethoprim (5.7 × 10⁻⁴ s⁻¹) > caffeine (1.3 × 10⁻⁴ s⁻¹) [18]. The degradation of these four micropollutants was accelerated in UVC/ClO₂, compared to direct ClO₂ oxidation or UVC photolysis. Ye et al. [56] reported that 95% flumequine was degraded (k_obs = 2.7 × 10⁻³ s⁻¹) in UVC/ClO₂ AOP, and its degradation rate gradually increased with ClO₂ dosage and UV intensity, but decreased as pH ascended. Though the combination of UVC and ClO₂ enhanced the micropollutant abatement via generating more reactive species (e.g., ^•OH and chlorine radical (Cl^•)), they are less effective than the well-documented UVC/H₂O₂, UVC/Cl₂, and UVC/NH₂Cl AOPs under the same initial oxidant dosages. For example, Tian et al. [57] compared the combination of UVC with different oxidants (i.e., Cl₂, NH₂Cl, ClO₂, and H₂O₂) in the degradation of iopamidol and reported the k_obs following the order of UVC/Cl₂ (1.9 × 10⁻² s⁻¹) > UVC/H₂O₂ (1.3 × 10⁻² s⁻¹) > UVC/NH₂Cl (1.0 × 10⁻² s⁻¹) > UVC/ClO₂ (4.4 × 10⁻³ s⁻¹) under the same conditions.

A novel UVC/ClO₂⁻ AOP was proposed to remove both ClO₂⁻ residue and micropollutants in water. UV photolysis after ClO₂ disinfection can effectively eliminate both ClO₂⁻ and carbamazepine by ^•OH and reactive chlorine species (RCS) generated from UVC/ClO₂⁻ [58]. The ^•OH generated from UVC/ ClO₂⁻ (Equations (19)–(21)) reacts with not only carbamazepine, but also ClO₂⁻ to generate ClO₂ (Equation (22)), which subsequently is activated by UV radiation to produce RCS (i.e., Cl^• and chlorine oxide radicals (ClO^•)). As the products of UVC/ClO₂⁻, Cl⁻ and ClO₃⁻ presented the decreasing and increasing yield, respectively, with the increasing ClO₂⁻ dosage [58]. Furthermore, Su et al., [59] developed a combined ClO₂⁻ photocatalysis technique for the degradation and detoxification of norfloxacin by dosing ClO₂⁻ in a visible light photocatalytic system. The degradation rate of norfloxacin in the combined ClO₂⁻ photocatalysis system was faster than the single photocatalytic system or the single chlorite system under irradiation. The regenerated ClO₂⁻ can be retransformed into ClO₂ based on the ClO₂⁻/ClO₂ dynamic interchange mechanism [59].

photocatalysts → photocatalysts (h⁺ + e⁻)

(19)

O₂ + e⁻ → O₂^•−

(20)

O₂^•− + H₂O → ^•OH

(21)

^•OH + ClO₂⁻ → ClO₂ + OH⁻    k = 6.3 × 10⁹ M⁻¹ s⁻¹

(22)

However, UVC/ClO₂ AOP suffer from several drawbacks: (1) low absorption of ClO₂ in the UVC range with the molar absorption coefficients of 60.7 M⁻¹ cm⁻¹ [19]; (2) high energy demand from UVC irradiation and low energy efficiency of low-pressure mercury ultraviolet (LPUV) lamps [19]; and (3) more emitted photons of UVC irradiation absorbed by background matrix components [27]. To address these issues, an emerging AOP combining longwave ultraviolet light (UVA) with ClO₂ (UVA/ClO₂) was proposed as an alternative to UVC/ClO₂ AOP because of the high molar absorptivity for ClO₂ at UVA wavelengths (ε359 nm = 1250 M⁻¹ cm⁻¹), reduced photon absorption by background matrix components at 365 nm, and high energy efficiency of UVA-LEDs. Recently, a novel UVA/ClO² AOP was proposed as an alternative to UVC/H₂O₂ AOP [27]. Furthermore, the novel UVA/ClO₂ AOP was conducted for the degradation of 19 micropollutants with the degradation percentages of 7 to 100% and the corresponding k_obs of 3.8 × 10⁻⁴–12.9 s⁻¹ (Table 3) [19]. They also suggested that compared to UVC/Cl₂, UVA/ClO₂ AOP produced similar or higher levels of reactive species at similar oxidant dosages, required lower energy input, and formed lower Cl-DBPs.

4.2. Degradation Pathways

The associated radical chemistry of UV photolysis of ClO₂ is rather complicated, as demonstrated in Equations (23)–(32). Studies have reported that ClO₂ has strong absorption bands in the near-ultraviolet region, and photoexcitation of ClO₂ can lead to the breaking of the O−ClO bond by two active product channels. ClO^• and oxygen atoms (O(³P)) [18] or excited state oxygen (¹O₂) [56] were proposed as the primary photo-fragments formed through the O−ClO bond breakage (Equation (23)). As for another channel, Cl^• and dissolved oxygen (O₂) were also observed from ClO₂ photolysis (Equation (24)) [27]. The generated product radicals ClO^•, O(³P)/¹O₂, and Cl^• from ClO₂ photolysis can undergo distinct chain reactions to generate secondary reactive species. ClO^• reacts rapidly with H₂O/HO⁻ to yield free chlorine (HOCl/OCl⁻) (Equations (25) and (26)). O(³P) reacts rapidly with O₂ to produce ozone (O₃) (Equation (27)). The reactions of Cl^• with Cl⁻, H₂O/HO⁻, or HOCl/OCl⁻ can form dichlorine radical anions (Cl₂^•−) (Equation (28)), HO^• (Equations (29) and (30)), or ClO^• (Equations (31) and (32)) [18].

ClO₂ + hv → ClO^• + O(³P) / 2ClO₂ + hv → 2ClO^• + ¹O₂

(23)

ClO₂ + hv → ClOO → Cl^• + O₂

(24)

2ClO^• + H₂O → HOCl + HClO₂     k = 2.5 × 10⁹ M⁻¹ s⁻¹

(25)

2ClO^• + HO⁻ → OCl⁻ + HClO₂    k = 2.5 × 10⁹ M⁻¹ s⁻¹

(26)

O(³P) + O₂ → O₃    k = 4 × 10⁹ M⁻¹ s⁻¹

(27)

Cl^• + Cl⁻ → Cl₂^•−    k = 8.5 × 10⁹ M⁻¹ s⁻¹

(28)

Cl^• + HO⁻ → ClOH^•−    k = 1.8 × 10¹⁰ M⁻¹ s⁻¹

(29)

ClOH^•− → HO^• + Cl⁻    k = 6.1 × 10⁹ M⁻¹ s⁻¹

(30)

Cl^• + HOCl → ClO^• + H⁺ + Cl⁻    k = 3.0 × 10⁹ M⁻¹ s⁻¹

(31)

Cl^• + OCl⁻ → ClO^• + Cl⁻    k = 8.3 × 10⁹ M⁻¹ s⁻¹

(32)

The degradation pathways of micropollutants in UV-activated ClO₂ oxidation include direct ClO₂ oxidation as discussed in Section 3.3, direct UV photolysis, ozonation, ^•OH-involved reactions, and RCS-involved reactions (i.e., Cl^•, ClO^• and Cl₂^•−) (Figure 1). Kong et al. [18] investigated the degradation pathways of four micropollutants of trimethoprim, iopromide, caffeine, and ciprofloxacin, with diverse chemical characteristics (i.e., caffeine bears electron-deficient moieties; trimethoprim and ciprofloxacin bear electron-rich moieties; iopromide bears photolabile moieties), during UVC/ClO₂. The degradation of caffeine was mainly caused by Cl^• (66.5%) and ^•OH (33.5%), whereas the degradation of trimethoprim, iopromide, and ciprofloxacin were mainly contributed by ClO₂ oxidation (72.2%), UVC photolysis (87.1%), in situ formed free chlorine (84.3%), respectively (

Table 4. The contribution of reactive species for micropollutant abatement in UV/ClO₂ process.

Compounds	Contribution of Reactive Species	References
UVC-LPUV
Trimethoprim	ClO2 oxidation (72.2%) •OH (11.5%) Cl• (8.9%) Other reactive species a (7.5%)	[18]
Iopromide	UVC photolysis (87.1%) •OH (2.0%) Cl• (5.4%) Other reactive species a (5.5%)
Caffeine	Cl• (66.5%) •OH (33.5%)
Ciprofloxacin	ClO2 oxidation (6.9%) UVC photolysis (8.0%) Cl• (0.3%) •OH (0.5%) in-situ formed free chlorine (84.3%)
UVC-LPUV
Flumequine	UVC photolysis (11.37%) 1O2 (14.72%) •OH (19.79%) RCS b (54.12%)	[56]
UVA-LEDs
Micropollutants c	ClO• (∼10−13 M) Cl• (∼10−15 M) •OH (∼10−14 M) Ozone (∼10−7 M)	[19]

^a other reactive species: O(³P), ClO^•, O₃, Cl₂^•−, and/or in-situ formed chlorine. ^b RCS: Cl^•, ClO^• and Cl₂^•−. ^c 19 micropollutants: iopromide, trimethoprim, caffeine, 17α-ethynylestradiol, 17β-estradiol, estrone, diclofenac, sulfamethoxazole, gemfibrozil, naproxen, ofloxacin, roxithromycin, carbamazepine, metoprolol, atenolol, metronidazole, bezafibrate, clofibric acid, ibuprofenreported.

). The degradation of flumequine in UVC/ClO₂ was contributed by 11.37% UV photolysis, 14.72% ¹O₂, 19.79% ^•OH, and 54.12% RCS (i.e., Cl^•, ClO^• and Cl₂^•−) (Table 4) [56]. The reaction pathways for the major species in UVA/ClO₂ AOP was recently well-summarized by Chuang et al. [27], which generates secondary reactive species such as ^•OH, Cl₂^•−, and O₃ with relatively high and stable concentrations. Additionally, the contribution of reactive species on the removal of 19 micropollutants followed an order of O₃ > ClO^• > HO^• > Cl^• and their concentrations were ∼10⁻⁷, ∼10⁻¹³, ∼10⁻¹⁴, and ∼10⁻¹⁵ M, respectively, in UVA/ClO₂ at a ClO₂ dosage of 5 mg L⁻¹ and a UV fluence of 47.5 mJ cm⁻² (Table 4) [19].

5. Research Gap and Future Research

Compared to UVC/ClO₂ AOP, UVA/ClO₂ AOP is practically promising for micropollutant abatement due to high absorption coefficients of ClO₂ in the UVA range, reduced photon absorption by the background matrix components, and high energy efficiency of UVA-LEDs. However, up until now, reports regarding degradation efficiency, degradation products, degradation pathways, reactive species, influencing factors (e.g., ClO₂ concentration, UV intensity, pH, and water matrices), and DBP formation are still limited during micropollutant abatement by UVA/ClO₂. Although the formation of halogenated DBPs during ClO₂-based oxidation is limited, inorganic products (i.e., ClO₂⁻, ClO₃⁻, and Cl⁻) might be a new concern when this technology is used in micropollutant abatement in water. Studies reported that the inorganic products were comprised of 50 to 70% of ClO₂⁻ and 30 to 50% of ClO₃⁻ and Cl⁻ in direct ClO₂ oxidation. However, research regarding the yield of ClO₂⁻, ClO₃⁻, and Cl⁻ in activated ClO₂ oxidation and influencing factors (e.g., ClO₂ dosage, pH, activation methods, and water matrices) on inorganic products in both direct ClO₂ oxidation and activated ClO₂ oxidation remain unclear. Recently, studies have reported that carbamazepine and norfloxacin were degraded by ^•OH and RCS generated from a novel UVC/ClO₂⁻ system, providing a possibility of “killing two birds with one stone”: eliminating both ClO₂⁻ residue and micropollutants. However, up until now, degradation efficiency, degradation products, degradation pathways, influencing factors (e.g., ClO₂⁻ concentration, UV intensity, pH, and water matrixes), reactive species, formation of organic or inorganic DBPs, and ClO₂⁻/ClO₂ dynamic interchange are still limited.