Degradation of Hydroxychloroquine from Aqueous Solutions Under Fenton-Assisted Electron Beam Treatment

Abstract

Challenges in the treatment and removal of recalcitrant emerging organic pollutants in wastewater prompt the development of advanced oxidative processes (AOPs). Hydroxyl radicals are non-specific and capable of reacting with a diverse range of pollutants of emerging concern. In this study, hydroxychloroquine (HCQ) was removed from aqueous solutions with removal efficiencies between 80 and 90%. The presence of H₂O₂, humic acid, and other inorganic ions negatively influenced the degradation efficiency. However, the presence of S₂O₈²⁻ was found to increase the removal efficiency, which was attributed to the formation of SO₄^•− in addition to •OH radicals. Additionally, Fenton-assisted electron beam treatment showed an improved removal of 2.88 × 10⁻⁴ M of HCQ with an average improvement of ≈10% at doses between 0.5 to 2.0 kGy in addition to the total organic carbon and chemical oxygen demand reduction. The H₂O₂ concentration and molar ratio of H₂O₂: Fe²⁺ influenced the removal capacity of the Fenton-assisted electron beam process. A degradation mechanism for HCQ has been proposed based on the reactions of •OH radicals and e_aq⁻.

1. Introduction

The future of wastewater decontamination is progressively deviating from the passive sequestration of pollutants, commonly practiced in conventional treatment plants, to the active and intentional destruction of pollutants in advanced wastewater treatment technologies. These initiatives are perpetuated by the realization that most pollutants are recalcitrant, persistent, degradation-resistant, bioaccumulative, toxic, and inadequately managed by current technologies. Additionally, conventional wastewater treatment methods sustain a phase transfer process that results in secondary pollution in the sludge, biosolids, manure, membrane filters, and used adsorbents [1]. Constructed wetlands (CWs) prominently used in wastewater treatment remove more than 50% of pharmaceuticals and personal care products (PPCPs). However, macrophytic biomass recovered from CWs and repurposed in pulp and paper industries or biofuel production still contains these pharmaceuticals [2]. Furthermore, the use of chemicals for decontamination is of environmental concern. Biological, chemical, physical, and physicochemical wastewater treatment has been investigated in addition to methods for the detection of low-concentration pollutants [3,4,5]. Alternative advanced oxidative processes play a crucial role in the development of sustainable and efficient wastewater treatment practices. These processes rely on the generation of highly reactive radicals that initiate the destruction of organic pollutants present in wastewater.

The risk associated with pharmaceutically active compounds and their metabolites in wastewater cannot be underscored. Pharmaceuticals released from industry, hospitals, and domestic sewage have been detected in wastewater pre- and post-treatment. Hydroxychloroquine (HCQ) is a synthetic quinoline derivative with anti-malarial and anti-inflammatory properties and is extensively used in oncology, rheumatology, and dermatology. Quinoline and its derivatives are associated with persistence, toxicity, carcinogenicity, and teratogenic properties. HCQ is primarily excreted via the renal route with 23–25% of it in unmodified form alongside its metabolites [6]. HCQ causes hemotoxicity, oxidative damage, and histopathological alterations in Catfish (Clarias gariepinus) [7], with other notable effects in aquatic wildlife [8]. HCQ has three nitrogen atoms in its chemical structure, with pKa values of 4.0, 8.3, and 9.7. However, only high values can be protonated, affecting its permeability in cell membranes’ intracellular distribution [9]. HCQ is lipophilic (Kow value: 3.84) with water solubility up to 26.1 mg/L and tends to effectively bind to tissues and is, therefore, highly distributed in blood and plasma [10]. Its water solubility presents a great risk for the species that inhabit the aquatic ecosystems [3]. Therefore, the removal and fate of HCQ and its metabolites as model pharmaceutically active compounds (PhACs) are of interest in wastewater treatment. Absorption technologies, photodegradation, and advanced oxidative processes including gamma and EB irradiation have been studied to remove HCQ [11].

Ionizing radiation effectively degrades a variety of recalcitrant organic pollutants, such as antibiotics, polycyclic aromatic hydrocarbons, heterocyclic aromatic hydrocarbons, and phenols in water and wastewater. Electron beam (EB) technology is a novel AOP with the added advantages of no secondary pollution or demand for additional chemicals. EB technology has been investigated for the degradation of recalcitrant organic pollutants in water and wastewater [12,13]. EB technology has been capable of reducing the chemical oxygen demand (COD) of wastewater before reverse osmosis (RO) membrane separation [13]. The gamma irradiation treatment of 100 mg/L HCQ aqueous solutions achieved a TOC removal efficiency of 98.5% after an 8 kGy absorbed dose, synonymous to complete HCQ mineralization at a dose rate of 26.31 Gy min⁻¹, and pH  =  6.2. A low initial pollutant concentration, higher dose rate, and neutral pH favored the process [14]. Additionally, gamma irradiation achieved complete degradation of 20 mg/L HCQ at a 1 kGy dose with complete mineralization [15]. Kabasa et. al. [11] studied the dependence of HCQ removal on the initial concentration, the pH of the aqueous solution, and the absorbed dose under EB treatment. In this study, EB irradiation was used in the degradation of aqueous solutions containing HCQ. Additionally, the EB coupled with Fenton oxidation was investigated to improve the efficiency of HCQ removal, reduce energy consumption compared with EB treatment [11], and reduce chemical reagent consumption compared with the Fenton process. To the best of the authors’ knowledge, there are not many studies on the EB–Fenton process [16,17].

2. Materials and Methods

2.1. Materials

Hydroxychloroquine sulfate powder (>99%), hydrogen peroxide (30%), ferrous sulfate heptahydrate (>99%), potassium dichromate (99%), methanol HPLC grade (>99.9%), sodium hydroxide (99%), silver nitrate (99%), acetonitrile HPLC grade (>99.9%), dibasic sodium phosphate (98%), sodium nitrate (99%), humic acid sodium salt, sodium persulfate (98%), orthophosphoric acid (98%), triethylamine (99.5%), sulfuric acid (95%), sodium hydrogen carbonate (99.5%), sodium carbonate (99.5%), and perchloric acid (95%) were purchased from Sigma Aldrich (Hamburg, Germany). Fresh solutions of HCQ and other solutions were prepared in distilled water from the Thermo-Fischer distillation unit from Merk (Darmstadt, Germany).

2.2. Analytical Techniques

A Jasco V670 spectrophotometer from Jasco (Krakow, Poland) alongside a VIS II spectrophotometer from Macherey Nagel (Dueren, Germany) together with a Shimadzu class VP HPLC with a DAD detector was used for the detection of HCQ with maximum absorption at 343 nm. The mobile phase of the HPLC method consisted of 70% acetonitrile and 30% 1.4 g/L dibasic sodium phosphate with 0.4% triethyl amine at pH 3.0. The column was a Gemini NX 5U (Phenomenex, Tianjing, China) c19 110A 250 × 4.6 mm operated at an oven temperature of 40 °C and flow rate of 1 mL/min. LC-MS was performed on an Agilent LC-MS (Agilent Technologies Co., Ltd., Beijing, China) in positive mode electron spray ionization. The mobile phase consisted of deionized water with 0.05% trifluoroacetic acid (phase A) and acetonitrile and 0.05% trifluoroacetic acid (phase B). The gradient washing procedure was 20% solvent B (0–0.2 min), 20 to 50% B (0.2–1.5 min), 50–90% B (1.5–1.6 min), 90 to 100% B (1.6–2.0 min), 100% B (2.0–2.5 min), 100 to 20% B (2.5–2.6 min), and 20% B (2.6–3.5 min). The flow rate was set at 0.5 mL/min, and the wavelength was 220 nm.

Different parameters were used to assess the degradation of HCQ. Nanocolor test kits purchased from Aqua Tests (Warsaw, Poland) were used for photometric determination of total Kjeldhal nitrogen (1.0–16 mg/L), total nitrogen (0.5–50 mg/L), nitrate (NO₃⁻), NH_4⁺ (0.2–8 mg/L), Cl⁻ (0.5–50 mg/L), total organic carbon (20–300 mg/L), organic acids (30–3000 mg/L), and chemical oxygen demand (50–300 mg/L) using the Nanocolor VIS II spectrophotometer (AQUA LAB, Warsaw, Poland). The CX-461 multimeter from Elmetron (Zabrze, Poland) was used to measure the pH and temperature of the solution before and after irradiation, and the dissolved oxygen was measured using a DO meter from Mettler Toledo (Warsaw, Poland).

2.3. Radiation Processing

Irradiation was performed using a batch system under the ILU6 accelerator, which is located at the Institute of Nuclear Chemistry and Technology, Warsaw, Poland. Irradiation was carried out with a beam energy of 1.65 MeV, a pulse frequency of 2 Hz, and a current of 50 mA. The absorbed dose was determined by the pulse frequency and irradiation time. The absorbed dose was measured using 0.0005 M potassium dichromate solution with 0.001 M silver dichromate, and all solutions were prepared in 0.1 M perchloric acid. Alanine solution and alanine pellet dosimeters were also used to assess the absorbed doses between 0.5 kGy and 7 kGy. Low-density polyethylene (LDPE) sleeve bags were filled with 35 mL of solution to obtain an approximately 1.5 cm depth of liquid and were placed 30 cm below the accelerator window for irradiation. Fenton-assisted EB processing was achieved by preparing different Fe²⁺: H₂O₂ molar ratios and adding them to the HCQ solution before irradiation. Control experiments for the H₂O₂ and Fe²⁺ solutions were also made to determine the effect of each Fenton reagent. All experiments were performed in ambient conditions (STP) without adjusting pH unless otherwise stated.

3. Results and Discussion

3.1. Degradation of HCQ Under EB Irradiation

Free radicals generated in the aqueous environment interact with HCQ and control its stability, degradation, and fate in the environment [6]. Alternatively, radicals generated in wastewater treatment processes propagate the destruction and removal of HCQ from wastewater. The interaction of high-energy EB radiation with water leads to the generation of ionized water molecules, excited water molecules, and presolvated electrons according to the following expression:

$$H_{2}O ⇝H_2{O}^{•+}, H_2{O}^{*}, e^{-}$$

Subsequent physicochemical and chemical reactions lead to the production of reactive radical species alongside ionized molecules and stable products. The corresponding radiation chemical yields (G-values) are provided in parenthesis in units of molecules/100 eV in the expression below.

$$H_{2}O ⇝H_2{O}_2\left(2.6\right), {}^{•}OH\left(2.7\right), H_3{O}^{+}, H^{•}\left(0.6\right), e_{aq}^{-}\left(2.7\right)$$

The reactions of the hydrated electrons ($e_{aq}^{-}$) and hydroxyl radical (•OH) with HCQ in aqueous solution are the key reactions leading to its degradation [18]. Similar conclusions were made in gamma irradiation studies [14,15].

3.1.1. Dose Effects on the Degradation of HCQ

A pulse radiolysis study showed that ^•OH radicals are the primary species reacting with HCQ, whereas the hydrated electron (e_aq⁻) played a minor role according to the rate constants in Equations (1) and (2). Rath et al. [19] proposed the formation of transient species in reactions of HCQ with the •OH radicals, with the [HCQ⁺] cation and the [HCQ: OH] adduct with the cation decaying slower than the OH adduct.

$$C_{18}{H}_{27}{N}_{3}OCl+e_{aq}^{-}⟶ C_{18}{H}_{27}{N}_{3}O+{Cl}^{-} 2.0\times {10}^9 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(1)

$$C_{18}{H}_{27}{N}_{3}OCl+{}^{•}OH⟶ C_{18}{H}_{26}{N}_{3}OCl+H_{2}O 9.5\times {10}^9 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(2)

Zaouak et al. suggested the •OH attack led to the dealkylation of the aromatic ring and breaking of the C-N bond in the aliphatic tertiary amine and subsequent formation of 7-chloro-4-quinolinamine and 1-(N-ethyl-N-hydroxy-methylene-amino)-4-aminopentane [14]. Eventually, ${}^{•}OH$ attacks 7-chloro-4-quinolinamine and breaks the C–Cl bond leading to the formation of 4-Amino-7-hydroxy-benzo pyridine accompanied by the release of chloride ions [15,19]. The reactions of the e_aq⁻ and •OH radicals drive the dechlorination process leading to the release of Cl⁻. The initial concentration of HCQ photometrically observed at 343 nm reduced with a corresponding increase in the absorbed dose from 0 to 2 kGy (Figure 1). The overall decrease in HCQ was 75% of the initial concentration at 2 kGy. The degradation of HCQ under EB irradiation was further evaluated under different irradiation conditions. Similar results were obtained in a previous study, with the efficiency increasing with a decreasing initial concentration of HCQ [11].

3.1.2. HCQ Degradation Under •OH, e_aq⁻, H•, and Aerated Conditions

The solutions of HCQ were subjected to different solvent conditions that allowed the selective generation of reactive species. The solutions were aerated by bubbling the solution with air (N₂(80%)-O₂(20%)) to increase dissolved O₂ concentration. In water radiolysis, O₂ effectively scavenges e_aq⁻ and H• radicals resulting in a reduction in e_aq⁻ and H• yields and an increase in less powerful and slowly reacting HO₂^• and O₂^•− yields (Equations (3) and (4)). Although the presence of these radicals enhances degradation, the presence of dissolved O₂ also suppresses degradation [20]. Although the presence of these radicals enhances degradation, the presence of dissolved O₂ also suppresses degradation [20]. The removal efficiency in aerated and Ar-saturated solutions where the dissolved oxygen was <1 mg/L was not significantly different (Figure 2).

H• + O₂ → HO₂^•   k = 1.27 × 10¹⁰ M⁻¹ .s⁻¹

(3)

e_aq⁻ + O₂ → O₂^•− + H⁺ → HO₂^• k = 1.84 × 10¹⁰ M⁻¹ .s⁻¹

(4)

The selective generation of •OH radicals was achieved in solutions saturated with N₂O gas. N₂O scavenges e_aq⁻ to form •OH radicals (Equation (5)). The highest removal of efficiency for HCQ under EB irradiation was in conditions favoring •OH generation (Figure 2).

H₂O+ e_aq⁻ +N₂O → N₂ + OH⁻ + •OH

(5)

The contribution of the e_aq⁻ was studied in Ar gas-saturated solutions containing 0.5 M tert-butanol at pH > 2. Tert-butanol scavenges the •OH (Equation (6)) and H• radicals, and all reactions were assumed to be exclusively from e_aq⁻.

•OH + (CH₃)₃COH → H₂O + (CH₃)₂^•(CH₂)COH

(6)

Under these conditions, the removal of HCQ was lower than in other conditions investigated for doses between 0.5 and 4.0 kGy (Figure 2). This is attributed to the rate constant for reactions of e_aq⁻ with HCQ, which are comparatively much lower compared to those of the •OH. The reaction of H• was studied in Ar-purged solutions with 0.5 M tert-butanol at a pH < 2 where the e_aq⁻ is converted into H• by reactions with H⁺ (Equation (7)).

e_aq⁻ + H⁺ → H•

(7)

The removal of HCQ is observed to be intermediate between that of the •OH radicals and the e_aq⁻ (Figure 2). Therefore, the removal of HCQ under EB irradiation is influenced by different solvent conditions that favor the selective generation of radicals. Conditions that favor the generation of •OH radicals give the highest removal efficiency for HCQ in aqueous solutions compared to e_aq⁻. In addition to irradiation conditions, the presence of several inorganic ions as well as organic constituents of natural water are known to influence the action of irradiation in treating aqueous solutions.

3.1.3. Effects of Inorganic Ions, H₂O₂, and Humic Acid

Both the carbonate and bicarbonate ions scavenge water radiolysis products leading to the formation of the carbonate radical anion that is less reactive toward organic pollutants according to Equation (8) and Equation (9), respectively. The HCO₃⁻ is a more efficient scavenger of •OH radicals compared to e_aq⁻ (k = 1.0 × 10⁶ M^−ls⁻¹), while the CO₃²⁻ is a better scavenger of the •OH radicals [21]. The HCQ removal efficiency decreased with increasing concentrations of CO₃²⁻ and HCO₃⁻ (Figure 3).

$${CO}_3^{2-}+{}^{•}OH\to {CO}_3^{•-}+{OH}^{-} \mathrm{k}=3.8\times {10}^8 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(8)

$${HCO}_3^{-}+{}^{•}OH\to {CO}_3^{•-}+H_{2}O \mathrm{k}=8.5\times {10}^6 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(9)

In natural waters, nitrates and nitrites are produced from the degradation of nitrogen-containing organic material. The determination of organic nitrogen (total Kjeldhal nitrogen) is an important process in water treatment, as nitrogenous waste poses a substantial threat to aquatic ecosystems. Additionally, the presence of nitrogen species also affects water treatment. NO₃⁻ ions are reduced to NO₂⁻ by scavenging e_aq⁻. NO₂⁻ reacts with the •OH and promotes the addition of NO₂ to aromatic solutes that may lead to the formation of nitrogen disinfection byproducts that are a challenge in wastewater treatment [22,23]. Additionally, NO₂⁻ scavenges e_aq⁻ less effectively than the •OH [24,25]. Additionally, NO₂⁻ scavenges e_aq⁻ less effectively than the •OH [24,25]. The NO₃⁻ anion reacts with both the e_aq⁻ (Equation (10)) and the H atom (Equation (11)), but the reaction with the latter is slower [26].

$${NO}_3^{-}+e_{aq}^{-}\to {NO}_3^{2•-} \mathrm{k}=9.7\times {10}^9 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(10)

$${NO}_3^{-}+H^{•}\to {HNO}_3^{•-} \mathrm{k}=1.4\times {10}^6 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(11)

The degradation of HCQ declined with an increasing NO₃⁻ concentration at a 2 kGy dose (Figure 3). NO₃⁻ ions at very low concentrations inhibit the action of water radiolysis products and have been used in the laboratory to shield H₂O₂ from attack by e_aq⁻, which is a major reaction leading to the generation of •OH radicals.

Persulfate activation is important in the generation of strongly oxidizing SO₄^•− radicals (E(SO₄^•−/SO₄²⁻) = 2.43 V). Different methods are experimentally used to achieve this activation. In the current experiment, EB irradiation generated SO₄^•− radicals at different concentrations of Na₂S₂O₈, and the effect on HCQ degradation was investigated. The persulfate reacts with e_aq⁻ to generate SO₄^•− radicals (Equation (12)). Additionally, reactions of SO₄^•− lead to further production of •OH radicals that would enhance removal efficiency (Equations (13) and (14)).

$${S_{2}O}_8^{2-}+e_{aq}^{-}\to 2{SO}_4^{•-}$$

(12)

$${SO}_4^{•-}+H_{2}O \to {}^{•}OH+H^{+}+{SO}_4^{2-} \mathrm{k}=6.5\times {10}^7 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(13)

$${SO}_4^{•-}+{OH}^{-}\to {}^{•}OH+{SO}_4^{2-} \mathrm{k}=6.5\times {10}^7 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1}$$

(14)

The removal of HCQ was higher in solutions containing persulfate and increased with an increasing concentration of persulfate (Figure 3). To investigate the contribution of SO₄^•− radicals on HCQ degradation, solutions containing 0.5 M tert-butanol were used. Solutions with 0.5 M tert-butanol and 2 mM persulfate (tBuOH/ SO₄^•−) where it is assumed SO₄^•− radicals are the predominant species had higher removal efficiencies (Figure S1). The removal efficiency was higher compared to solutions containing 2 mM persulfate and also solutions where reactions of the e_aq⁻ dominated (tBuOH). Overall, solutions with persulfate had higher removal for HCQ compared to the EB process.

H₂O₂ and Fe²⁺ are commonly used Fenton reagents to generate •OH. In radiolysis, H₂O₂ is produced and is a precursor for •OH radicals. H₂O₂ decomposes in the presence of cations present in wastewater to further generate •OH radicals. Solutions of HCQ were mixed with either H₂O₂ or ferrous sulfate at concentrations ranging from 0.1 to 2 mM and irradiated under an EB accelerator at doses ranging from 0.5 kGy to 7 kGy. The removal efficiency of the HCQ concentrations was observed to be lower in solutions containing either H₂O₂ or Fe²⁺ compared to the solutions without any additives (Figure 3). The reduction in removal efficiency was more pronounced with increasing concentrations of Fe²⁺. Radiolysis in the presence of excess H₂O₂ leads to the generation of the perhydroxyl radical (${HO}_2^{·}$), which is a weaker oxidant compared to the •OH radical. The increasing concentration of Fe²⁺ also increases the turbidity. Additionally, competition between these additives and the target pollutant (HCQ) for the reactive species impacts the overall efficiency of the process. Even though H₂O₂ is produced during the radiolysis of water and expected to react in Fenton/Fenton-like processes with Fe²⁺ present in the solution, this was not observed. The concentrations of H₂O₂ needed for the effective Fenton process are between 50 and 80 ppm [13]. However, other studies using Fe²⁺ concentrations between 0.1 and 0.6 mM indicate an enhanced degradation of a 20 mg/L solution of sulfamethazine by gamma irradiation at a dose of 1 kGy [27].

The presence of dissolved organic matter in wastewater influences wastewater treatment. In photodegradation studies, dissolved organic matter (DOM) was a possible source of •OH and a sink for •OH radicals [28]. Humic acid (HA) had synergistic effects on atrazine degradation under gamma irradiation [29]. In contrast, 0.1 mM of humic acid suppressed the degradation of diclofenac at 1 kGy gamma irradiation [30]. In the present study, humic acid with increasing concentration led to a decrease in the removal efficiency of HCQ under EB irradiation (Figure 3).

3.2. Degradation of HCQ by Fenton Oxidation

Fenton and Fenton-like processes are important for the generation of •OH radicals that are non-specific and highly reactive oxidants. The ratio of concentrations of H₂O₂:Fe²⁺ is important in sustaining the Fenton reaction and the eventual degradation of organic contaminants. Heterogenous catalysis is deemed advantageous over homogenous catalysis in overcoming the limitations of cost, limited operational pH, sludge generation, and catalyst recovery. Furthermore, Fenton oxidation provides high performance, non-toxicity, and simplicity with the convenience of operating at room temperature and atmospheric pressure to achieve the desired oxidation of organics [31]. In this experiment, the initial concentrations of H₂O₂ ranged from 0.1 to 2 mM, with H₂O₂:Fe²⁺ molar ratios adjusted to a range between 0.2 times and 20 times. The removal efficiency of HCQ improved with increasing H₂O₂ concentration and was highest at a 2 mM H₂O₂ concentration for all molar ratios (Figure 4).

Generally, in conditions with higher proportions of H₂O₂ compared to Fe²⁺, the removal efficiency was higher. This is mainly because Fe²⁺ is the catalyst, and H₂O₂ is the oxidant and dominant source of ^●OH radicals. When the 2 mM H₂O₂ concentration was 20 times (20×) the ferrous concentration, the highest removal efficiency of 78% was achieved after 3 hrs reaction time (Figure 4).

Insufficient H₂O₂ decreases removal efficiency due to a deficiency of ^●OH generated by H₂O₂ decomposition. However, excess H₂O₂ may scavenge the ^●OH radicals and reduce efficiency (Equation (15)). Additionally, the less powerful perhydroxyl radical reduces efficiency, but in some instances, it enhances efficiency.

$$●\mathrm{OH}+{\mathrm{H}}_2{\mathrm{O}}_2 ⟶ {\mathrm{H}}_2\mathrm{O}+{HO}_2^{·}$$

(15)

Conversely, H₂O₂ has negative implications on COD, BOD, and microorganisms present in biological treatment while inadvertently increasing cost. The important parameters pH, H₂O₂ dosage, catalyst dosage, and temperature affect Fenton and Fenton-like processes. The heterogeneous Fenton-like processes are found more effective at a pH of 3. However, neutral conditions (even alkaline conditions) could still achieve better treatment efficiency. These observations are mainly due to the different solubilities of a metal ion (such as the Fe ion) on the surface of a catalyst and the different activities of active sites on the catalyst surface. The pH 2.5–3.0 is regarded as optimum for the degradation of most organic compounds under homogeneous Fenton-like processes. At pH < 2.5, the scavenging effect of the ^●OH radicals by H⁺ (Equation (16)) becomes stronger [17,31]. At pH > 3.0 the hydrolysis and precipitation of Fe³⁺ in the solution can reduce the catalytic capacity [32]. This pH range has also been reported to favor the degradation of various organic chemicals of interest [11].

^●OH + H⁺ + e⁻ ⟶ H₂O

(16)

3.3. Degradation of HCQ by Fenton-Assisted EB Process

Fenton and ozonation processes are widely used AOPs in the treatment of industrial wastewater. The harsh operational conditions (e.g., pH of about 3) of the Fenton process limits its practical application. Moreover, large amounts of H₂O₂ and Fe²⁺ required in the Fenton process result in the generation of iron-containing sludge, which necessitates further treatment, removal, and disposal. Moreover, large amounts of H₂O₂ and Fe²⁺ required in the Fenton process result in the generation of iron-containing sludge, which necessitates further treatment, removal, and disposal. Similarly, in the ozonation process, the low solubility, low mass transfer of ozone, and high operation cost limit its practical application [13]. However, synergistic effects in the simultaneous applications of two or more AOPs show improved efficiency in the removal of pollutants. EB treatment (EB) is versatile and adaptive in incorporating other treatment processes to complement it. These can be completed either pretreatment, simultaneously, or post-EB. Pretreatment using an EB has been shown to enhance the biological treatment of refractory organic pollutants [33,34]. Fenton-like processes were chosen in the context of this study utilizing the generation of highly reactive •OH from the catalytic decomposition of H₂O₂ in the presence of Fe²⁺ (Equation (17)) alongside the •OH radicals generated during water radiolysis.

Fe²⁺ + H₂O₂ ⟶Fe³⁺ + OH⁻ + ^∙OH

(17)

$${\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2 ⟶ {\mathrm{Fe}}^{2+}+{HO}_2^{\bullet }+H^{+}$$

(18)

The slower back reaction (Equation (18)) limits the number of ferrous ions in the solution, and by utilizing the e_aq⁻ generated in water radiolysis, the conversion of Fe³⁺ to Fe²⁺ can be accelerated (Equation (19)).

$$e_{aq}^{-}+{Fe}^{3+}\to {Fe}^{2+}$$

(19)

Solutions of HCQ with selected molar ratios of Fenton reagents were subjected to EB irradiation. Each of the solutions contained H₂O₂ concentrations ranging from 0.1 to 2 mM and adjusted so that the concentration of the H₂O₂ was 0.2 to 20 times the concentration of the ferrous salt. The highest removal efficiency of ≈94% was obtained with a H₂O₂ concentration of 2 mM, which was 20 times (20×) the ferrous concentration at a 2 kGy applied dose (Figure 5).

The removal of HCQ under the Fenton-assisted EB process was faster compared to EB at doses between 0.5 and 2 kGy (≈10% improvement Figure S2). At higher ratios of H₂O₂:Fe²⁺ (5× to 20×), the removal efficiency increased with increasing H₂O₂ concentrations. However, at higher ratios of Fe²⁺ (0.5× to 0.2×), the removal efficiency decreased with the increasing ferrous sulfate concentrations. In high H₂O₂ concentrations, •OH radicals are more readily available from both the Fenton process and EB. On the other hand, increasing Fe²⁺ ratios caused scavenging and affected the turbidity of the solution.

It is asserted that, under Fenton-assisted EB, the Fe²⁺ consumed (Equation (17)) is regenerated following the reduction of Fe³⁺ according to Equation (19). These reactions sustain the Fenton process and lead to the generation of more •OH radicals. This enhances the removal efficiency compared to the conventional EB. Additionally, the self-quenching of reactive species, such as •OH and e_aq⁻, decreased due to the reaction between e_aq⁻ and H₂O₂ (Equation (20)). The e_aq⁻ reacts with H₂O₂ (~10¹⁰ M⁻¹s⁻¹) to produce •OH, which increases the concentration of •OH.

$$e_{aq}^{-}+H_2{O}_2\to •OH+{OH}^{-}$$

(20)

The electrochemical oxidation (EO) of HCQ with initial concentrations in the range of 36–250 mg/L using a boron-doped diamond (BDD) anode in aqueous solution (j = 20 mA/cm², initial pH = 7.1, T = 25 °C, 0.05 M Na₂SO₄) achieved 100% reduction in the initial HCQ concentration. The removal efficiency was found to be faster at lower HCQ concentrations [19]. However, the complete removal of 2.88 × 10⁻⁴ M (125 mg/L) HCQ was achieved after 300 min, which is a comparatively longer time compared to EB and Fenton-assisted EB process. In contrast to the Fenton-assisted EB process, electrochemical oxidation generated •OH at the anode through the electrolysis of a suitable electrolyte. Additionally, H₂O₂ is generated through the cathodic reduction of O₂, and the presence of Fe²⁺ further enhanced the generation of •OH radicals through Fenton reactions. Furthermore, the presence of Na₂SO₄ leads to the generation of S₂O₈²⁻ and SO₄^•− that have higher reduction potentials, as previously explained in Figure 3. Similar to the Fenton-assisted EB process, the removal efficiency of HCQ under EO increased with increasing concentrations of H₂O₂ and Fe²⁺ but was suppressed when the concentrations exceeded an optimum. Similar to the Fenton-assisted EB process, the removal efficiency of HCQ under EO increased with increasing concentrations of H₂O₂ and Fe²⁺ but was suppressed when the concentrations exceeded an optimum.

4. Degradation Evaluation of HCQ Under Fenton-Assisted EB Process

Physical/chemical changes in the irradiated samples allude to changes in the initial composition of the samples. Parameters such as pH, dissolved oxygen, chloride ions in solution, total nitrogen, total Kjeldhal nitrogen, nitrates, ammonium ions, total organic carbon, and chemical oxygen demand are used to assess the efficiency of the treatment process. Furthermore, these parameters are indicators of the degradation mechanism. Additionally, products of HCQ degradation have been assessed using the LC-MS technique, and a mechanism for the degradation of HCQ has been proposed.

4.1. Changes in pH During HCQ Removal

The pH of the solutions was evaluated before and after radiolysis. The solutions containing H₂O₂ had an initial pH of 5.5 before irradiation. The pH was reduced to a stable pH of 3.5 at the end of irradiation. Increasing the concentration of H₂O₂ from 0.1 to 2 mM did not alter the initial pH of the HCQ solution (Figure S3A). However, the ferrous sulfate solution's initial pH dropped with the ferrous salt concentration increasing from pH 5 at 0.1 mM to pH 3.6 at 2 mM. With EB irradiation, the pH at all concentrations of the ferrous sulfate went to a stable final pH of 3.5 (Figure S3B) similar to results under normal EB without either Fenton reagents [11]. The reduction in pH has been attributed to the formation of organic acids (Figure 6a). Similar changes in pH under electro-Fenton oxidation of HCQ solution have been attributed to the formation of oxalic and oxamic acids [19], while the degradation of diuron under ionizing radiation is attributed to the formation of ketones and aldehydes [35]. The radiation-induced degradation of HCQ under gamma irradiation led to the formation of several products including oxamic and oxalic acids [14]. Similar products have been identified in the electro-Fenton process [19]. The oxidative decomposition of byproducts through •OH radical attack opening aromatic rings leads to the production of carboxylic acids. The carboxylic acids are slowly oxidized and require the consumption of a high irradiation dose to be mineralized into carbon dioxide and water.

During the Fenton-assisted EB process, the pH of the solutions ranged between 3 and 3.5 upon the addition of the Fenton reagents and maintained this pH after irradiation. The change in pH is attributed to the ferrous salt (FeSO₄). Characteristically, Fenton oxidation is more favorable in acidic conditions at pH values ranging between 2 and 3. At pH > 5, insoluble Fe³⁺ precipitates increase turbidity while also competing for reactive radicals [17]. The solution pH has an influential role in the oxidation potential of the •OH [36]. At pH 0, E⁰ = 2.8 V, whereas at pH 14, E⁰ = 1.59 V, meaning that at low pH values, the reactions of ^•OH are dominant. Additionally, the rate of decomposition of H₂O₂ at pH >4 is low, leading to a lower ^•OH production, and H₂O₂ decomposes into H₂O and O₂. Furthermore, pH also affects the formation of water radiolysis products. An acidic pH favors the formation of H• atoms, whereas strongly basic conditions favor the generation of e_aq⁻ from reactions of the ^•OH and OH⁻. Therefore, these conditions will influence the reactions leading to the degradation of target pollutants.

4.2. Cl⁻ Generation

Dehalogenation and degradation technology for highly toxic substances is an important strategy in water purification [37]. Studies focus on dehalogenation, but no effective technique has been described [38]. Chemical dehalogenation methods convert the halogenated compounds to less toxic dehalogenated compounds, in contrast to incineration for the disposal of such compounds [39]. Dissociative electron attachment, as shown in Equation (1), is accompanied by the release of chloride ions (Cl⁻). Under Fenton-assisted EB (EB-F) treatment, the total Cl⁻ ion release at 2 kGy was higher compared to the EB process (Figure 6b). Similar dechlorination has been reported while using gamma radiation [14].

Dissociative electron attachment is a common reaction of the e_aq⁻ with halogenated compounds. The evaluated rate constants for this reaction with HCQ have been provided in Equation (1) [18]. The chlorine group is presumed to be responsible for the toxicity of chlorinated organic compounds. The dechlorination of HCQ, therefore, implies a decrease in the toxicity of the aqueous solution [40]. The rate of dechlorination under EB-F was greater than for EB for all the doses. This imparts superior capability to the Fenton-assisted EB-supported process. Other than reactions involving the e_aq⁻, dechlorination occurred under the action of the •OH radical (Figure 6b). The direct substitution of a Cl atom by OH is believed to be a key reaction in this process [35]. In studies on the dechlorination of chlorobenzene, it was observed that highly substituted benzenes had a higher rate of degradation and dehalogenation compared to singly substituted benzenes [38].

4.3. Nitrogen

Nitrogen-containing organic wastewater is a challenge in wastewater treatment. The nitrogen atom incorporated in the cyclic ring of quinolone increases hydrophilicity, good solubility, and low biodegradation [6,41]. About 75% nitrification was achieved during the EB radiolysis of a solution of HCQ at an applied dose of 7 kGy (

Table 1. Variation in NO₃⁻, NH_4⁺, TKN, TOC, and COD with increasing dose under EB and EB-F processes.

Dose (kGy)	NO3− (mg/L)		NH4+ (mg/L)		TKN (mg/L)		TOC (mg/L)		COD (mg/L)
	EB	EB-F	EB	EB-F	EB	EB-F	EB	EB-F	EB	EB-F
0	0	0	0	0	11.08	11.9	73.50	73.50	149.00	149.00
0.5	5.99	0.82	0.24	2.04	5.70	6.64	69.75	64.00	145.67	112.00
1	6.79	8.36	0.42	2.66	5.17	3.76	69	62.33	144.67	107.67
2	7.68	8.4	0.65	2.84	4.04	4.16	66.25	59.88	136	106.67
4	8.16	8.88	0.65	2.81	3.36	4.16	68.9	59.60	136.67	105.67
7	8.48	9.22	1.05	2.79	2.65	3.66	68.67	59.42	136.67	103.33

TKN—total Kjeldahl nitrogen; TOC—total organic carbon; COD—chemical oxygen demand.

). The total Kjeldahl nitrogen (TKN) was reduced with a simultaneous increase in NO₃⁻ with an increasing applied irradiation dose as expected in the successful nitrification of nitrogen-containing organic wastewater.

Organic nitrogen present in wastewater is converted into inorganic nitrogen (nitrate, nitrite, and ammonium) at the end of the treatment process. In aerobic conditions, ammonium is utilized as a nutrient by microorganisms and further oxidized into nitrite and eventually into nitrate. The concentration of NH_4⁺ in the solution increased with the increasing irradiation dose (Table 1). The formation of NH_4⁺ was faster and higher under the Fenton-assisted EB process compared to the EB process.

4.4. Chemical Oxygen Demand

Dissolved oxygen (DO) represents the oxygen dissolved in water and is available to living aquatic organisms. The dissolved oxygen is consumed as organic matter decays. In high levels of organic pollution, BOD, or COD, a large amount of DO is consumed through aerobic microorganisms to decompose the organic matter, which causes a reduction in the DO level. The dissolved oxygen during the Fenton-assisted EB (EB-F) degradation of the solution of HCQ showed a slight decrease (Figure S4). The concentration of dissolved oxygen increases with increasing molar ratios of Fenton reagents. Several reactions (Equations (21)–(24)) were proposed for this increase in molecular oxygen in Fenton reactions [20,42].

HO₂^• + H₂O₂ → O₂ + H₂O + •OH

(21)

Fe³⁺ + HO₂^• → Fe²⁺ + O₂ + H⁺

(22)

HO₂^• + •OH → O₂ + H₂O

(23)

2HO₂^• → O₂ + H₂O₂

(24)

The complete mineralization of target pollutants was evaluated using the oxygen demand. The biochemical oxygen demand (BOD), chemical oxygen demand (COD), and total organic carbon (TOC) measure oxygen demand in polluted water. The TOC and COD slightly decreased with increasing absorbed doses during the EB process (Table 1). This reduction was further improved by the EB-F treatment. The decrease in the COD and TOC during the irradiation process is indicative of the transformation of the initial compound into other organic degradation products such as organic acids (Figure 6a). These products are less susceptible to oxidation by initial water radiolysis products and require additional processing time to decompose.

Similar TOC elimination was reported under the electrochemical oxidation of HCQ [19]. The lower removal efficiency was attributed to the production of carboxylic acids and aliphatic chains that were more refractory than the initial HCQ. The observed radiolytic decomposition of 100 mg/L HCQ under gamma irradiation increased COD and TOC elimination with an increasing absorbed dose with the removal efficiency >90% at doses up to 8 kGy [14]. Complete mineralization of 20 mg/L HCQ was achieved under gamma irradiation at 3 kGy with no intermediates being detected [15]. The level of mineralization was dependent on the initial pollutant concentration, possibly due to the generation of higher degradation intermediate concentrations that scavenge the reactive radicals. A similar dependence of removal efficiency on HCQ concentration was previously reported [11]. Additionally, the contact time is an important variable in the treatment process. A longer contact time results in higher TOC and COD reduction as in the case of gamma radiolysis. Zaouak et al. (2022) also demonstrated that higher dose rates gave higher removal efficiency under gamma irradiation owing to reduced self-combination of reactive species [14]. However, EB has comparatively higher dose rates compared to gamma irradiation. The mineralization under gamma irradiation can be ascribed to the longer contact time.

4.5. Degradation Products of HCQ Degradation Under EB Irradiation

LC-MS analysis was performed to determine the degradation products following the EB irradiation of HCQ. The [M+H]⁺ peak for HCQ was observed at m/z = 336 with mass fragments at m/z = 292, m/z = 247, and m/z = 74 (Figure S5A). Other additional mass fragments were identified and recorded in a similar analysis [14,15,43]. In this study, the degradation products of HCQ were identified at m/z 352, 175, 308, 334, and 370 as shown in

Table 2. m/z values of proposed degradation products based on LC-MS results.

m/z	Chemical Formula	Fragment	Retention Time (min)
336	C18H26ClN3O	[M+H]+	1.41
370	C18H26ClN3O3	M + 2OH	1.1
352	C18H27ClN3O2	M + OH	8.39
334	C18H26N3O3	M + 2OH − Cl	1.36
308	C16H21ClN3O	M − CH2CH3	1.87
175	C₉H21N2O	M − C₉H8NCl	1.11

M = C₁₈H₂₆ClN₃O.

(Figure S5A–F).

The formation of new peaks with increasing irradiation from 0.5 kGy to 2 kGy is shown from the chromatograms in Figure S6 and the decrease in the MS intensity of HCQ with an increasing absorbed dose and the formation of degradation products (Figure S7). Photodegradation products of HCQ have previously been reported as N-dehydroxyethyl-7-dechloro-7-hydroxy HCQ (m/z 274), dechlorinated HCQ (m/z 302), N-dealkylated HCQ (m/z 179), HCQ N-oxide (m/z 352) HCQ, and desethylhydroxychloroquine (m/z 308) [44]. Additionally, under gamma irradiation, degradation products such as HCQ N-oxide (m/z 352), m/z 174, and m/z 262 were reported [15]. Furthermore, products with m/z 174, m/z 179, and m/z 161 have been reported also under gamma radiolysis of aqueous solutions of HCQ [14]. Bensalah et al. obtained m/z = 179 and 175 following α dealkylation of HCQ under electrochemical oxidation.

The irradiation of HCQ in water or isopropanol identified desethyl-chloroquine (m/z 292), desethanolhydroxychloroquine (m/z 308), and a third product, which is HCQ without the N-C₄H₁₀O at the tertiary amine [45]. Chonker et al. also identified compounds m/z 292, m/z 308, m/z = 263, and 339 as metabolites of HCQ. Three degradation products m/z 292, m/z 324, and m/z = 364 with lower retention times than HCQ have also been reported in photodegradation studies on HCQ. The formation of compounds m/z 292 and m/z 324 detected during irradiation experiments result from the shortening of the side chain on the nitrogen atom, by the loss of ethene from HCQ to yield desethylhydroxychloroquine, or by primary double shortening of the side chain on the nitrogen [46].

The Ecological Structure Activity Relation software v2.2 was used to predict the acute and chronic toxicity of HCQ and its degradation products on fish, daphnia, and green algae based on the EU-Directive 93/67/EEC [47]. The compounds are classified as highly toxic (LC₅₀ and EC₅₀ < 1 mg/L), toxic (LC₅₀ and EC₅₀ 1 to 10 mg/L), harmful (LC₅₀ and EC₅₀ 10–100 mg/L), and harmless (LC₅₀ and EC₅₀ > 100 mg/L). HCQ (m/z 336) was highly toxic to fish, daphnia, and green algae when considering chronic toxicity but was toxic to daphnia and green algae when considering acute toxicity (Figure 7). The degradation products with m/z 179, 175, and 334 were predicted to be harmless to fish but were classified as harmful to daphnia and green algae. Except for m/z 292, most of the degradation products of HCQ were less toxic than the HCQ parent with respect to acute and chronic toxicity.

4.6. Degradation Mechanism for HCQ

Mathematical abstractions propose multiple sites of the •OH attack on the HCQ molecule. However, the •OH addition to the aromatic ring is thermodynamically favored based on Gibbs free energy [48]. Furthermore, the addition reactions of •OH to both the benzene and pyridine rings of quinoline are energetically favorable with reaction energies ranging from 18.6 to 23.9 kcal/mol to form OH adducts in water [49]. The formation of similar OH adducts has been reported in the gamma radiolysis of HCQ [14].

The degradation of HCQ is proposed to proceed via the OH addition to the aromatic ring that leads to the formation of m/z 370, 352, 279, and 334, as shown in the degradation scheme in Figure 8. Additionally, α dealkylation of the aromatic ring leads to the formation of m/z 179 and 175. N-dealkylation at the tertiary nitrogen is likely to lead to the formation of several products with m/z 308, 292, and 279. Dehalogenation is additionally a prominent reaction of most halogenated organic species, as previously discussed [40]. Dechlorination occurs through dissociative electron attachment reactions involving the e_aq⁻ or direct substitution of the Cl atom by the OH in addition to reactions involving the aromatic ring [35].

The main products that were detected in the current study are m/z 370, 292, 308, 352, 175, and 334 alongside the inorganic nitrates, ammonium, and chloride species. All organic by-products are proposed to undergo oxidative decomposition through the •OH radical attack opening aromatic rings to form carboxylic acids and releasing inorganic nitrogen species predominantly in the form of NO₃⁻ and NH₄⁺ [14]. The carboxylic acids are then slowly mineralized to CO₂ [19].

5. Conclusions

The degradation of 2.88 × 10⁻⁴ M (125 mg/L) HCQ under EB irradiation achieved >70% degradation at 2 kGy. The synergistic application of Fenton reagents provided additional •OH that improved efficiency by 18%, from 76% to 94%. Compared with the sole Fenton process, the consumption of H₂O₂ (2 mM) and Fe²⁺ (0.1 mM) was less, and the removal efficiency of HCQ was improved by 18% compared with the sole EB irradiation. Furthermore, the self-quenching of reactive species, such as •OH and e_aq⁻, was decreased, while additionally, the e_aq⁻ accelerated the reduction of Fe³⁺ into Fe²⁺. The release of Cl⁻, ${\mathrm{N}\mathrm{O}}_3^{-}$, and ${\mathrm{N}\mathrm{H}}_4^{+}$ was better under EB-F compared to EB. Additionally, COD and TOC reduction were higher in EB-F.

EB processing is faster than gamma [14,15] or Fenton processing, which usually takes longer, an hour or more. In this work, we investigated higher concentrations of HCQ (125 mg/L) using EB irradiation. It is very well known that the lower the initial concentration of the pollutant (as low as 20 ppm [15]), the higher the removal efficiency. The EB process is capable of performing both the degradation of organic compounds as well as the disinfection of the irradiated volume without the use of chemical additives, which is desired in wastewater treatment processes. Additionally, the EB is capable of handling copious volumes of wastewater, i.e., 1000 m³ of dye effluents [50], 10,000 m³/day of wastewater [51], and 645 m³/day of sludge [52]. Therefore, other than the initial cost of the installation of the facility, EB is promising in the treatment of effluents. Some studies postulated the cost of EB treatment to be more cost-effective than conventional methods, including certain AOPs [53]. However, a complex wastewater matrix poses a challenge in efficiency as the components compete for reactive species. EB can be integrated alongside the Fenton process for better efficiency as has been shown in the present study.