Degradation of hydroxychloroquine in aqueous solutions under electron beam treatment

HCQ sulfate powder (MW = 433.95, 99%), NaOH (99%), potassium dichromate (99%), silver nitrate (99%), perchloric acid (95%), and H2SO4 (98%) were purchased from Sigma-Aldrich (Merck-Germany). Fresh solutions of HCQ and other solutions were prepared in distilled water from the Thermo Fisher distillation unit from Merck (Germany).

A Jasco V670 spectrophotometer (Poland) was used for the detection of HCQ (>1 mg·L^-1) with the maximum absorption at 343 nm. Nanocolor test kits from Mercherey Nagel (Germany) purchased from Aqua Lab (Poland) tests were used for the photometric determination of total Kjeldhal nitrogen (1.0–16 mg·L^-1), total nitrogen (0.5–50 mg·L^-1), nitrate (NO3−), NH4+ (0.2–8 mg·L^-1), Cl^- (0.5–50 mg·L^-1), total organic carbon (TOC) (20–300 mg·L^-1), and chemical oxygen demand (COD) (50–300 mg·L^-1). Photometric tests were performed using the Nanocol-or VIS II spectrophotometer from Mercherey Nagel (Germany). The pH measurements were done using an Elmetron CX-461 multimeter (Poland) designed for accurate measurements of pH. Dissolved oxygen (DO) was measured using a Mettler Toledo DO meter purchased from Sigma-Aldrich (Poland).

Irradiation was performed using a batch system with aqueous solutions containing varying concentrations of HCQ in distilled water. Irradiation was performed on the ILU6 accelerator at the energy of 1.65 MeV, 2 Hz, and 50 mA. A solution of 0.0005 M potassium dichromate mixed with silver nitrate solution in 0.1 M perchloric acid was used for dosimetry. Additionally, alanine dosimeters were used to determine the applied radiation doses delivered to the aqueous system [24–27]. Low-density polyethylene (LDPE) sleeve bags were filled with 35 mL of HCQ solution and irradiated under the accelerator window. Different initial pH values were selected to study the effect of pH on the degradation efficiency of solutions of HCQ under electron beam processing. The pH values were attained by adjusting the pH of the solution with 0.1 M NaOH and 0.1 M sulfuric acid.

The interaction of radiation with water molecules leads to ionization and excitation events that eventually through physical-chemical and chemical reactions lead to the formation of reactive radical species among other molecules with corresponding radiation chemical yields (G-values in parenthesis) in molecules/100 eV. H2O⇝H2O+,H2O*,e−→•OH(2.8),eaq−(2.8),H·(0.6),H2O2(0.7),H3O+(2.6),H2(0.45)

The hydroxyl radical is a strongly oxidizing species synonymous with the hydroxy radicals generated in advanced oxidation processes. Additionally, reducing species such as the hydrated electron and hydrogen atom are simultaneously produced in the radiolysis of water. These highly reactive species in the natural aqueous environment interact with HCQ and control its stability, degradation, and fate [7, 8]. Alternatively, these radicals can be generated in wastewater-treatment processes to propagate the destruction and removal of HCQ from wastewater. Using electron beam processing, the reactions of hydrated electrons (eaq−) and the hydroxyl radical (^•OH) with HCQ in an aqueous solution lead to its degradation. Figure 1a shows the stepwise absorption réduction observed for 2.88 × 10-4 M of HCQ solution at 343 nm and 330 nm with increasing dose from 0 kGy to 7 kGy.

Fig. 1.

The overall decrease in HCQ concentration was 82% of the initial concentration (Fig. 1b). This observation is consistent with observations in pulse radiolysis where the concentrations of the reactive species increase with increasing dose, therefore, increasing their contribution to the degradation of HCQ with increasing dose [28, 29]. Rath et al. [30] using electron pulse radiolysis found that the degradation of 1 × 10^-4 M HCQ was faster in reactions with ^•OH radicals compared to the hydrated/ aqueous electron. However, observations regarding other aminoquinoline derivatives such as chloroquine and amodiaquine show that reactions with hydrated electrons are faster (Table 2).

The corresponding reactions for these reductive and oxidative reactions with HCQ are provided in Eq. (1) and Eq. (2) [33]. 1C18H27 N3OCl+eaq−→C18H27 N3O+Cl− 2C18H27 N3OCl+·OH→C18H27 N3O+H2O

The reaction between HCQ and ^•OH radicals (Eq. (2)) forms transient intermediate species having two absorptions at 330–340 nm (Fig. 1a). Rath et al. suggested the formation of both [HCQ⁺] cation and [HCQ:OH] adduct at the same reaction rates. However, the cation decayed at a slower rate compared to the adduct [33]. Additionally, under gamma radiolytic degradation, ^•OH attacks led to the dealkylation of the aromatic part through the breaking up of the C–N bond in the aliphatic tertiary amine to form 7-chloro-4-quinolinamine and 1-(Nethyl-N-hydroxy-methyleneamino)-4-amino pentane as the main byproducts [22]. Subsequently, 4-amino-7-hydroxy-benzo pyridine is formed after a hydroxyl radical attack onto 7-chloro-4-quinolinamine breaking the C–Cl bond and releasing chloride ions. The presence of antioxidants like ascorbic acid and gallic acid slowed down the degradation. This implied that the stability of HCQ could be influenced in an oxidative environment in the presence of these two compounds [33].

The influence of initial pollutant concentration on the radiolytic degradation efficiency of HCQ was studied. The increase in pollutant concentration negatively affected the degradation efficiency. The degradation efficiency decreased with increasing HCQ concentration as shown in Fig. 2a. Similar results have been obtained in the radiolytic decomposition of other organic compounds as well as the degradation of HCQ under gamma irradiation [22].

Fig. 2.

From the experimental results and under normal conditions, it was possible to achieve 90% degradation of 25 mg·L^-1 HCQ and about 76% for 125 mg·L^-1 for a maximum dose of 7 kGy. Most radiolytic decompositions of the target pollutants can be described by the pseudo-first-order kinetic reaction [22, 23]. 3−ln(CC0)=kD

where C and C0 represent the final and initial concentrations of the target pollutant at an absorbed dose D, respectively. The rate constant also called the dose k constant describes the degradation rate per kGy. The reaction rate decreases with the increase in the concentration of HCQ (Fig. 2b). Similar observations were derived from a plot of the decay rates vs. dose in Fig. 2c and the corresponding R² values in Table 3. These observations can be attributed to the scavenging of the reactive radiolysis products by the target pollutant and also by the intermediates [34]. Similar first-order kinetics have been reported in the hydroxyl radical-propagated degradation of HCQ under electrochemical oxidation [14]. Additionally, the rate of removal of HCQ decreased with increasing concentration of HCQ.

The solvent pH influences the proportion of the different radicals generated during the radiolysis of the water. Under alkaline conditions, ^•OH readily reacts with OH-to generate O^•-, which is a less powerful oxidant thereby reducing the concentration of ^•OH and the degradation efficiency (Eq. (7)) [35–37], However, more hydrated electrons (eaq−) are generated in alkaline conditions (Eq. (6)). Similarly, in acidic media, the eaq− react with H+ to produce H^• radicals that are less reactive toward most pollutants (Eq. (4)). Neutral or acidic media are preferable for the degradation of HCQ under gamma irradiation [22, 23]. 4eaq−+H+−H· 5·OH+H·→H2O 6H·+OH−→eaq−+H2O 7·OH+OH−→O·−+H2O

Different pH values were chosen for the start of the irradiation process. From Fig. 3a, >80% of the initial HCQ concentration was removed at a maximum dose of 7 kGy for pH between 2 and 10. However, at a pH of 12, the removal efficiency declined (-60%). In experiments utilizing electrochemical oxidation with boron-doped diamond anodes, about 60% removal efficiencies were observed at similar pH ranges [14]. Similar to these studies, an acidic pH of 2 favored the decomposition of HCQ but de-composition reduced with increasing pH. However, for the electron beam process, the degradation efficiency was also better at pH >7. Additionally, for pH between 4 and 8, the pH was observed to decrease with increasing dose and stabilized at a pH of about 3.5 (Fig. 3b). Similar observations were made under gamma irradiation at pH between 4 and 8 [23]. In the present study, pH of 2 and 8 gave the highest removal efficiency ≥84%.

Fig. 3.

The solution at an initial pH of 2 was at a stable value throughout the irradiation process. The solutions at initial pH of 10 and 12 dropped to values of 6.5 and 11.5, respectively. The decrease in the pH alludes to the formation of acidic or less basic intermediates during the degradation of HCQ. Additionally, the pH influences the molecular properties of HCQ, which has three functional groups with pKa values of <4.0, 8.3, and 9.7. At acid and neutral conditions, two of the functional groups exist in protonated forms [30, 38]. This may facilitate the rupture of C–N bonds by ^•OH radicals attack and lead to the release of the branched group [14]. HCQ is a basic substance, completely protonated at acidic pH values, as H2HCQ²⁺. At neutral values of pH, two protonated forms of HCQ are formed: H2HCQ²⁺ and HHCQ⁺ [39]. Similarly, the degradation of HCQ was high at pH values of 8 and 10. In alkaline media, the ratio between the protonated form HHCQ⁺:HCQ is 1:6, meaning that the HCQ is mostly deprotonated. Higher degradation in alkaline solution indicates that deprotonation increases the electron density on HCQ and favors the electrophilic attack by reactive oxygen species, such as hydroxyl radicals. The quinoline ring is more susceptible to the attack of hydroxyl radicals at pH 9 than at pH 4 [7]. Regarding the half-life times (t1/2), the values increase from a slightly acidic pH to a neutral one and then decrease once again. During gamma irradiation, the favorable range for pH values was from slightly acidic to neutral [40]. However, the HCQ elimination percentage as a function of the irradiation dose was found higher at pH 6.2 and pH 10 than at acidic pH for experimental conditions: [HCQ] = 100 ppm, dose rate = 26.31 Gy·min^-1 gamma irradiation [22].

The increasing concentration of solvated electrons at high pH leads to a decrease in the concentration of hydroxyl radicals so the redox reaction is extremely rapid. At pH = 6.2, all reactive species are free and are not involved in other reactions. Gamma radiolysis of 20 ppm HCQ at pH of 4, 6.8, and 9 showed an overlap in removal efficiency in acidic to neutral conditions [23]. A similar observation is made for pH between 4 and 7 under EB treatment (Fig. 3a). However, in basic conditions, the removal efficiency increased with dose. Under electrochemical oxidation, >60% of the initial HCQ concentration was eliminated for pH between 2 and 12 similar to what is reported in this work. However, the efficiency decreased with increasing pH to 12 [14]. The decrease in the removal efficiency at pH 12 could be attributed to reactions in Eq. (7). The pH of wastewater is a vital component before, during, and after treatment for the eventual discharge of wastewater.

The degradation byproducts discussed in this section are based on the degradation of 125 mg·L^-1 of HCQ solution under electron beam irradiation at neutral pH.

Changes in solution pH

The pH of the solution was observed to change from slightly acidic of pH 6.5 before irradiation to an acidic pH of 3.5 at the end of irradiation (Fig. 4). At the lower pH, degradation of HCQ was diminished according to Figs. 1 and 2. Similar changes in pH during the degradation of organic compounds have been attributed to the formation of lower molecular--weight carboxylic acids such as oxamic and oxalic acid, ketones, aldehydes, and ions [14, 41].

Fig. 4.

Cl^- generation

From Eq. (1), dissociative electron attachment is accompanied by the release of chloride ions (Cl^-). There was a total Cl^- ion release at 7 kGy (Fig. 5). Similar dechlorination has been reported while using gamma radiation [22]. This is synonymous with dissociative electron attachment, which is a common reaction of the hydrated electron with halogenated compounds. The evaluated rate constants for this reaction have been provided [33]. The chlorine group is presumed to be responsible for the toxicity of organic compounds. The dechlorination of HCQ treatment, therefore, implies a decrease in the toxicity of the aqueous solution. However, in the present study, toxicity assessment was not performed.

Fig. 5.

Nitrogen

Nitrogen-containing organic wastewater poses a challenge in wastewater treatment as it is repeatedly transformed in the nitrogen cycle and enters water and wastewater via agricultural, domestic, and manufacturing wastes. The nitrogen atom in the cyclic ring of quinolone increases the hydrophilicity, solubility, and low biodegradation, which increases the potential for the incidence of HCQ in water environments [7, 8]. Nitrogen in the form of organic nitrogen and ammonia in freshly polluted water are converted by a biochemical process into ammonium to be utilized as a nutrient by microorganisms in the treatment process. Under aerobic conditions, the organic nitrogen is converted into ammonia and then further oxidized into nitrite and eventually into nitrate. The total Kjeldahl nitrogen (TKN) monitors the degree of contamination of the discharged water. Figure 6a shows that about 75% of the nitrification was achieved during the electron beam radiolysis of a 2.88 × 10-4 M solution of HCQ at 7 kGy dose. A reduction in TKN with a simultaneous increase in NO3− was also observed as is expected in the successful nitrification of nitrogen-containing organic wastewater. The formation of NH4+ was also recorded (Fig. 6b).

Fig. 6.

Chemical oxygen demand

DO is the amount of oxygen dissolved in water and is available to living aquatic organisms indirectly inferring to water quality. The DO is consumed as organic matter decays. In high levels of organic pollution, biochemical oxygen demand (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 DO during the radiolysis of 2.88 × 10^-4 M of HCQ solution showed a slight decrease from 4.5 mg·L^-1 to 3.8 mg·L^-1 as shown in Fig. 7.

Fig. 7.

In addition to the DO, another important parameter is the oxygen demand. The complete mineralization of target pollutants is the main aim of wastewater treatment. However, most processes achieve minimal to partial mineralization. Mineralization can be evaluated using the oxygen demand, which determines the amount of organic pollution in water, waste loadings of treatment plants, and the efficiency of treatment processes. Industries that produce vaccines/antitoxins generate wastewater containing very high BOD, COD, total solids, colloidal solids, toxicity, and odor. Figure 8 shows the COD and TOC variation with increasing doses during electron beam treatment of a 2.88 × 10-4 M solution of HCQ. There was a slight decrease in the COD, but TOC remained fairly constant during the irradiation process. This is indicated by the transformation of the initial compound into other organic degradation products such as organic acids. These products are less susceptible to oxidation by initial water radiolysis products.

Fig. 8.

In studies on the radiolytic decomposition of HCQ under gamma irradiation, COD and TOC elimination were observed to increase with increasing applied dose [22]. The removal efficiency was >98.5% at doses up to 8 kGy. Higher dose rates, low pollutant concentration, and neutral pH facilitated mineralization. Similar TOC elimination efficiencies have been reported under electrochemical oxidation [14]. Electron beam treatment has comparatively higher dose rates than gamma irradiation but did not show higher TOC reduction. The slower removal efficiency can be attributed to the production of carboxylic acids and aliphatic chains that were more refractory than the initial HCQ.