Industrial and domestic discharge of active pharmaceutical ingredients into the environment is inadequately regulated worldwide. The implications of these compounds, when present in water, soil, and air, to both plants and animals cannot be overlooked [1–3]. Repurposing wastewater effluents to serve as a source of irrigation water for the cultivation of food crops poses insurmountable risks. Antibiotics negatively affect the growth of food crops and affect the soil biota [4–7]. They tend to accumulate in plant roots and shoots and pose a threat to the food chain [8]. Similarly, pharmaceutical-degradation products and metabolites due to their lipophilic property, high toxicity, and resistance to degradation are of great concern [9, 10]. Wastewater can be classified by the presence of pathogens and physicochemical constituents that pose a risk to human and environmental health. Economical, efficient, and environmentally friendly treatment technolo-gies to remove these constituents are imperative in addressing these concerns [11, 12]. Moreover, detecting these contaminants in the ambient environment is paramount [13–15]. These compounds classified as emerging pollutants (EPs) are directly injected from industries, institutions, and sewage effluents. Electron beam technology is efficient for wastewater treatment and addresses the shortfalls of conventional wastewater treatment. It has shown efficacy for a myriad of emerging organic compounds of pollution concern.
Chloroquine (CQ) and hydroxychloroquine (HCQ) piqued interest as proposed therapeutics for severe acute respiratory syndrome coronavirus 2 (
Previous works on the degradation and removal of CQ from aqueous solutions mainly involved photostability studies, advanced oxidative processes (AOP) with or without catalysts, and adsorption technologies. Catalytic processes involving ferrate--Fe(VI) show CQ aromatic ring dealkylation and chloride ion substitution under the influence of pH, temperature,
Methods previously used for the removal of CQ from aqueous solutions
Methods | Conditions | Removal efficiency | Ref. |
---|---|---|---|
Membranes | |||
Membrane bioreactors – tyrosinase enzyme on |
pH 7.5, 20 h | 98% with 140 ± 6 mg·g-1 No apparent capacity loss over three consecutive cycles | [33] |
Adsorbents | |||
Activated carbon Palm kernel ( |
Large surface areas, strong mechanical characteristics | [27] | |
PPAC-ZnO | 313 K 10 ppm CQ | 78.89% Adsorption capacity increases with temperature | [28] |
A-GO hydrogel | Adsorption 63 mg·g-1 | [29, 30] | |
GAC-GO | Equilibrium time 18 h | 37.65 mg·g-1 Adsorption | [32] |
Organo-clay raw kaolinite treated with citric acid | 20 mg·L-1 CQ 120 min | 99.28% Maximum sorption capacity is 4.03 mg·g-1 | |
Soybean hull residues functionalized with iron oxide nanoparticles (SBH-Fe3O4) | 120 min 318 K | Adsorption capacity 98.84 mg·g-1 Reuse five cycles | [34] |
Iron and magnesium comodified rape straw biochar (Fe/Mg-RSB) | pH (3–11) CQ 4–25 mg·L-1 at 180 r·min-1 for 8 h 308 K | Adsorption capacity of 42.93 mg·g-1 | [35] |
MOF sheet, namely BUC-21(Fe) FeSO4·7H2O, 1,3-dibenzyl-2-imidazoli-done-4,5-dicarboxylic acid (H2L) and 4,4’-bipyridine (bpy) | pH = 5.0 30 min | 100% |
[36] |
Catalysts | |||
Ferrate-Fe(VI) | CQ 10 μM Fe(VI) 40–180 mM time 1–20 min | 59% CQ removal Algae, antimicrobial, toxicity reduction | [19] |
CWAO HEO – (MgCuMnCoFe) Ox | Oxygen pressure of 15 bar, catalyst dosage of 1.4 g·L-1, and temperature of 230°C | 34.6% and 41.2% higher than that without the HEO system | [11] |
Single cobalt atoms in a defined Co–N3 coordination structure | pH range (3–11) employing the SA Co-N-C (30) | 100% | [24, 25] |
Biochar-supported RM-BC activated persulfate process | 20 mg·L-1 40 min | 84.8% | [26] |
Carbon nanotube-loaded CoFe2O4 (CoFe2O4@CNTs) composite | 10 mg·L-1 CQ pH 7 | Mineralization efficiency 33%, removal efficiency 98.7% | [37] |
Advanced oxidation processes | |||
SR-AOP | Peroxymonosulfate (PMS, |
P25M175-94.6% within 30 min | [23] |
UV/PS | pH = 6.9 10 min | 91.3% CQ reactions with •OH and |
[38] |
Photocatalysis-activated SR-AOP over PDINH/MIL-88A(Fe) composites | 10.0 mg·L-1 CQ P25M175 30 min | 94.6% Good reusability and stability | [23] |
Electrocoagulation | 66.89 mA·cm–2, 600 rpm 60 min electrolysis time 3 mg·L-1 CQ, pH = 6.5 | 95% dissolved aluminum electrodes 0.228 kg·m-3 energy consumption of 12.243 kWh·m-3 | [39] |
EFP | Carbon felt cathode and BDD anode | 92% (TOC) | [20] |
FBER | BDD electrodes batch recirculation mode 9 h, pH 5.38, 34.4 mA·cm-2, and liquid flow rate ( |
Degradation 89.3%, COD 51.6%, mineralization 53.1% energy consumption 0.041 kWh·L-1 | [40] |
Electro-Fenton with pyrite (FeS2)-modified graphite felt (FeS2/GF) cathode | pH of 3.0 FeS2 loading-10 mg, current density 150 mA, electrode spacing 2.0 cm | 83.3 ± 0.4% 60 min CQ removal, retains 60.0% CQ removal in consecutive batch tests | [41] |
A-GO, agar-graphene oxide; BDD, boron-doped diamond; CQ, chloroquine; CWAO, catalytic wet air oxidation; EFP, electro-Fenton processes; FBER, flow-by electrochemical reactor; GAC-GO, granular activated carbon–graphene oxide; HEO, high entropy oxide; PPAC-ZnO, plantain peel-activated carbon-supported zinc oxide; RM-BC, red mud catalyst; SR-AOP, sulfate radical-advanced oxidation process; TOC; total organic carbon; UV/PS, UV-activated persulfate system; PDINH/MIL, 34,910-pyrenetetracarboxydiimine (Materials of Institute Lavoisier).
CQ phosphate powder (98.5–100.0%), potassium dichromate (99%), silver nitrate (99%), sodium hydroxide (99%), sulfuric acid (95%), and perchloric acid (95%) were purchased from Sigma-Aldrich (Merck, Germany). Distilled water from a Thermo Fischer scientific distillation unit (Merck, Germany) was used in the preparation of all solutions for this study.
The Jasco-V670 UV-VIS (Poland) and Macherey Nagel Nanocolor VIS II (Germany) spectrophotometers were used for the detection of CQ (>1 mg·L-1) at the maximum absorption wavelength of 343 nm. Nanocolor test kits produced by Macherey Nagel (Germany) were purchased from Aqua Labs (Poland) for photometric determination of total Kjeldahl nitrogen (TKN) (1.0–16 mg·L-1), total nitrogen (0.5–50 mg·L-1), nitrate
Radiation processing was done in batch mode. The aqueous solutions containing different CQ concentrations were prepared in distilled water. Irradiation was performed on the ILU6 accelerator at an energy of 1.65 MeV, 2 Hz, and 50 mA. Potassium dichromate (0.0005 M) solution in perchloric acid with silver nitrate and alanine dosimeters were used to assess the applied doses ranging from 0.5 kGy to 7 kGy [43–46]. Low-density polyethylene (LDPE) sleeve bags were filled with 35 mL of CQ solution and irradiated under the accelerator window. The initial pH values of the CQ solution were adjusted with 0.1 M NaOH and 0.1 M H2SO4 to study the effect of pH on the degradation efficiency under electron beam processing.
The stepwise absorption reduction observed for CQ solution at 343 nm and at 330 nm with increasing radiation dose from 0 kGy to 7 kGy is illustrated in Fig. 1. Additionally, the corresponding reduction in CQ concentrations with increasing dose is shown in Fig. 2. The concentration of CQ was reduced from 125 mg·L-1 to 22 mg·L-1, which corresponds to an 82% reduction in the initial CQ concentration following electron beam irradiation. At the lowest absorbed dose achieved in this experiment (0.5 kGy), the initial CQ concentration of 125 mg·L-1 was reduced by ≈50%.
Free radicals generated during the radiolysis of water drive the reactions leading to the degradation of organic compounds in water. These radicals are generated through a combination of physical, physical-chemical, and chemical processes that occur within 10-12 s following the interactions of radiation in the water matrix. The species formed and their corresponding G-values are presented in Table 2.
Species formed in the radiolysis of water and their corresponding G-values and reaction rates (
Species | G (mmol·J-1) | Molecules/100 eV | |
---|---|---|---|
0.28 | 2.8 | 4.8 × 1010 | |
•OH | 0.28 | 2.8 | 7.3 × 109 |
•H | 0.06 | 0.6 | – |
H3O+ | 0.26 | 2.6 | – |
H2 | 0.045 | 0.45 | |
H2O2 | 0.07 | 0.7 |
The hydroxyl radical and hydrated electron play important roles in the degradation of CQ molecules via the addition reaction and dissociative electron attachment according to Eqs. (1) and (2), respectively. Reaction rates for the reactions with •OH and
Initial concentrations of CQ were varied to study the effect of CQ concentration on the degradation efficiency. The removal efficiency decreased with increasing CQ concentration at the same applied radiation dose (Fig. 3a). Generally, concentrations of CQ between 25 mg·L-1 and 125 mg·L-1 have degradation efficiencies ≥75% for a maximum radiation dose of 7 kGy. However, for lower concentrations (25 mg·L-1), more than 99% removal efficiency was achieved.
The degradation of CQ can be represented using a pseudo-first-order relation (Eq. (3)) [48].
where
Reaction rates (
Concentration CQ (mg·L-1) | ||
---|---|---|
75 | 1.6567 | 0.9935 |
100 | 1.3603 | 0.9982 |
125 | 1.1224 | 0.9891 |
The solvent pH influences the percentage of the different radicals generated during the radiolysis of the water. Under alkaline conditions, •OH readily reacts with OH- to generate O•- thereby reducing the concentration of •OH and hence the degradation efficiency.
In acidic media, the hydrated electrons
Therefore, under electron beam irradiation, the pH enhances or inhibits the production of reactive oxidizing or reducing species and therefore affects the degradation process. In this study, the degradation of CQ was observed to increase with increasing pH. The removal efficiency decreased slightly with increasing pH at irradiation doses between 0.5 kGy and 2 kGy. The acidic pH of 2 to neutral pH of 7 had relatively higher removal efficiencies for CQ (≈84%). In neutral to an alkaline pH of 10, the removal efficiency was at ≈80% at irradiation doses of 7 kGy (Fig. 4). However, slight increments in the removal efficiency were observed with increasing pH from neutral to pH = 10 at an irradiation dose of 3 kGy. According to Eq. (2), reactions with the hydrated electron have a higher reaction rate compared to hydroxyl radical reactions. More hydrated electrons are produced in alkaline conditions, which lead to higher degradation of CQ. However, in strongly alkaline conditions (pH = 12), the removal efficiency was inhibited. This could be attributed to the reactions of hydroxyl radicals with hydroxide (Eq. (4)). Conversely, at lower pH, hydrated electrons are consumed, which would result in lower removal efficiency for CQ. Additionally, hydroxyl radicals are known to have a higher reduction potential (2.72 V) at acidic pH compared to alkaline conditions (1.89 V); therefore, their role in CQ degradation would be affected at different pH [49].
The changes in the initial pH during the irradiation process are presented in Fig. 5. Other than affecting the production of water radiolysis products, the pH value can affect the reaction pathways by changing the structure and characteristics of contaminants such as their solubility. Certain compounds retain their molecular form when the pH is less than its pKa. However, when the pH >pKa, the compound loses a proton and becomes negatively charged. The pKa of CQ is 8.76, and it is a weak acid with a tendency to donate electrons. Therefore, at pH <pKa (between pH 2 and 7), CQ is in the non--ionic form (protonated). At pH >7, the divalent form of CQ
The results for the degradation byproducts are based on the degradation of 125 mg·L-1 of CQ solution under electron beam irradiation. The CQ solution was irradiated between 0.5 kGy and 7 kGy. The presence and degradation of organic compounds influence parameters used in wastewater such as pH, TOC, COD, chloride ion, and nitrogen. The formation or reduction in these physicochemical parameters was used to evaluate the effectiveness of CQ treatment under electron beam irradiation.
During the radiolysis of the CQ solutions, the pH was observed to reduce from 6.5 before irradiation to 3.2 at 7 kGy (Fig. 6). A similar decrease in pH has been observed in the degradation of ciprofloxacin where pH decreased from 7.00 to 4.22 at 5.0 kGy, consistent with the formation of lower molecular weight organic acids and dissolved ions resulting from EB treatment [51, 54].
In the degradation of CQ under electro-Fenton oxidation, the change in pH was attributed to the formation of carboxylic acids such as oxamic and oxalic acids [20]. Additionally, pH changes in the degradation of diuron under radiation were attributed to the formation of aldehydes and ketones [55].
During the radiolysis of organic compounds with good leaving groups such as halogenated compounds, dissociative electron attachment as shown in Eq. (2) is proposed to be the most preferred reaction process of the aqueous electron [56, 57]. The Clions present in CQ were continuously released with increasing radiation dose (Fig. 7). Almost complete Cl-detachment was observed at 7 kGy indicative of the dechlorination of the CQ compound.
Most polychlorinated cyclic compounds are stable, lipophilic, and tend to bioaccumulate in the environment with slow degradation. Chlorinated compounds i.e., pesticides, biphenyls, dibenzodioxins, and dibenzofurans are genotoxic (mutagenic and carcinogenic) and therefore pose a risk to living organisms [58–60]. Therefore, the dechlorination of CQ under electron beam treatment would be considered to reduce its toxicity. However, in the present study, no toxicity studies were undertaken to ascertain the reduction in toxicity.
The generation of
Organic nitrogen (TKN) is normally converted into inorganic nitrogen (
Dissolved oxygen (DO) reflects the amount of oxygen dissolved in water and is available to living aquatic organisms. The decay of organic matter consumes the DO and the presence of excess organic material, biochemical oxygen demand (BOD), or COD in lakes and rivers causes eutrophication (oxygen deficiency). A large amount of DO is consumed by aerobic microorganisms to decompose the organic matter. The DO is affected by the water temperature, salt concentration, atmospheric pressure, and other conditions, and the saturation level decreases with increases in water temperature. During the radiolysis processing of wastewater, the DO will react with radicals formed. This may significantly reduce O2 concentrations. However, from the experiment, the dissolved O2 varied from 5.5 mg·L-1 before irradiation to 4.7 mg·L-1 at 7 kGy as shown in Fig. 9.
High DO has been found to influence the mineralization of azo dyes under electron beam irradiation by improving the cleavage of the aromatic rings and promoting the reduction of TOC [63]. The DO was more important for the destruction of substituted aromatic rings in the dye molecule than the chromophore and needed to be maintained at optimal concentrations during electron beam radiolysis [64]. However, DO scavenges
The oxygen demand determines the amount of organic pollution in water, waste loadings of treatment plants, and in evaluating the efficiency of treatment processes. BOD, COD, and TOC are used to measure the oxygen demand. There was a slight decrease in the COD and TOC with increasing doses and the decreasing concentration of CQ (Fig. 10). These observations in TOC and COD degradation indicate the formation of other organic byproducts that are not easily degraded in the present conditions [66]. In other studies, even with the apparent insignificant reduction in COD and TOC, the resulting degradation products exhibited a higher biological oxygen demand and therefore were more susceptible to biological degradation.
CQ solution of 125 mg·L-1 concentration has been degraded with an efficiency of 82% under an electron beam process at doses ranging from 0.5 kGy to 7 kGy in ambient conditions. Both the hydroxyl radical and hydrated electron played roles in the decomposition of CQ molecules. The initial concentration of the CQ solution affected the degradation and the reaction rate; hence, the degradation efficiencies decreased with increasing CQ concentrations under electron beam treatment. The release of chloride ions was considered indicative of the dissociative electron-attachment process during the electron beam processing of the CQ solution. Dechlorination could imply a reduction in the toxicity of the resulting products. The pH of 125 mg·L-1 of the CQ solution was measured and observed to drop from the initial pH of 6 to a pH of 3.5 at the end of processing. This can be attributed to the formation of lower molecular weight carboxylic acids and ions as has been reported in other studies. It was observed that at acidic pH, the removal efficiency was higher than at neutral pH. However, alkaline pH exhibited increased removal efficiency. This could be attributed to the state of the CQ molecule at different pH (protonation). Additionally, the effects of pH during the radiolysis of water affect the radiolysis products (•OH and