Oxidation of chloroquine drug by ferrate: Kinetics, reaction mechanism and antibacterial activity

Graphical abstract

Abstract

Chloroquine (CLQ) is required to manufacture on a larger scale to combat COVID-19. The wastewater containing CLQ will be discharged into the natural water, which was resistant to environmental degradation. Herein, the degradation of CLQ by ferrate (Fe(VI)) was investigated, and the biodegradability of the oxidation products was examined to evaluate the potential application in natural water treatment. The reaction between CLQ and Fe(VI) was pH-dependent and followed second-order kinetics. The species-specific rate constant of protonated Fe(VI) species (HFeO₄⁻) was higher than that of the FeO₄²⁻ species. Moreover, increasing the reaction temperature could increase the degradation rate of CLQ. Besides, HCO₃⁻ had positive effect on CLQ removal, while HA had negative effect on CLQ removal. But the experiments shows Fe(VI) could be used as an efficient technique to degrade co-existing CLQ in natural waters. During the oxidation, Fe(VI) attack could lead to aromatic ring dealkylation and chloride ion substitution to form seven intermediate products by liquid chromatography-time-of-flight-mass spectrometry (LC-TOF-MS) determination. Finally, a pure culture test showed that the oxidation of CLQ by Fe(VI) could slightly increase the antimicrobial effect towards Escherichia coli (E.coli) and reduce the toxicity risk of intermediates. These findings might provide helpful information for the environmental elimination of CLQ.

1. Introduction

Malaria disease is the most widespread and essential parasitic disease in human [1], [2], [3]. Chloroquine (CLQ), as the primary antimalarial prevention drug, is widely administrated to treat all types of malarial infections [4]. Currently, CLQ has attracted increasing attention for COVID-19 treatment due to its low price and significant improvement of immune system [5], [6], [7]. It is required to manufacture this drug on a larger scale to combat COVID-19 that infected millions of people. Hence, a large amount of medical wastewater polluted by CLQ will be discharged into the environment. CLQ has been detected in aquatic environment at concentrations ranging from ng/L to μg/L. Olaitan et al. (2014) found that the median concentration of CLQ identified in different water samples from Nigeria was 2.12 μg/L [8]. However, like many antibiotics such as ciprofloxacin, erythromycin, penicillin and sulphisoxazole, CLQ has high potential to be persistent and bioaccumulative, hard to remove in waste [9]. Besides, due to the antiviral and antibacterial characteristics, CLQ can be transferred to biological organisms in a more toxic way and becomes a persistent emerging micropollutant in the futural surface water [10], [11]. The high risk of the pollution necessitates more attention to limit its harmful effects on human health and the environment, including ozone-depleting substances [12], [13], heavy metal bioaccumulation and persistence [14], [15]. Therefore, there is an urgent demand for development of technically viable, low-cost, and energy-saving alternatives for CLQ elimination in water.

The previous studies only focused on the chemical stability and detection methods of CLQ in water [16], [17], [18], few researches concerned about the oxidation and fate of CLQ. Nord et al. found CLQ was photochemical degraded on irradiation with 240–600 nm light in aqueous solutions (pH 7.4), leading to N-dealkylation [19]. Kanakapura et al. investigated the forced degradation of CLQ in water by acid, alkaline, peroxide, heat and light-induced. CLQ was degraded by 50% under alkaline hydrolysis conditions. Besides, Coelho et al. developed a degradation study submitting CLQ to alkaline hydrolysis and chemical oxidation with diluted H₂O₂ [9]. Recently, Sondos et al. used the electro-Fenton method to achieve high-efficiency degradation of CLQ in wastewater, which was as high as 92% under optimal conditions [11]. However, established drinking water treatment technologies cannot provide a sustainable solution to CLQ pollution. Thereby, it is necessary to find an efficient drinking water treatment technology to eliminate or reduce the harm of CLQ in surface water to respond to the huge usage of CLQ.

Ferrate (Fe(VI)) has been demonstrated for treating drinking water with the function of oxidation, flocculation, sterilization, disinfection and algae removal [20], [21], [22]. Fe(VI) is a strong oxidant with supercharged iron molecule, which has the redox potential of 2.20 V and 0.78 V in acidic and alkaline, respectively [23], [24]. Compared with the common oxidants (e.g., permanganate, ozone, and hypochlorite), Fe(VI) has recently been shown to react with electron-rich organic moieties including pharmaceuticals [25], [26], [27]. In addition, Fe(VI) is similar to permanganate, the reduction product of ferric ion will transform into ferric hydroxide, which can be used as an in-situ coagulant in the subsequent coagulation [28], [29], [30]. Fe(VI) is environmentally friendly for treating water without producing secondary pollution [31]. In recent years, researches on the oxidation of antibiotics by Fe(VI) have been forthcoming, including the reaction kinetics, but the knowledge of antimalarial drugs (e.g., CLQ) and products of the oxidations under natural water conditions are very limited [32], [33]. Significantly, due to the high toxicity and biological resistance of degradation products, the removal may not necessarily obviate the antibacterial properties completely [34], [35]. After oxidation of the parent drug molecule, the antibacterial activity of reaction products is also missing in the literature.

The objectives of this study were to: (1) investigate the oxidation of CLQ by Fe(VI) at different Fe(VI) concentrations and temperature; (2) determine the reactions of Fe(VI) and CLQ by studying the kinetics of oxidation (species-specific rate constants) as a function of pH; (3) identify the high molecular weight transformation byproducts of CLQ by Fe(VI) treatment and propose the potential degradation pathways; and (4) evaluate the antibacterial activity of the reaction mixtures against Escherichia coli (E.coli).

2. Materials and methods

2.1. Standards and reagents

CLQ and diammonium 2,2′-azinobis-(3-ethylbenzothiazoline-6-sulfonate) (ABTS) were obtained from Aladdin Industrial Corporation (Shanghai, China) with purity higher than 98%. K₂FeO₄ of high purity (>92%) was synthesized by a wet oxidation method in the laboratory [36]. The Fe(VI) stock solutions were freshly prepared within 30 min prior to their use. Other reagents at least analytical purity including disodium hydrogen phosphate (Na₂HPO₄), sodium dihydrogen phosphate (NaH₂PO₄), sodium tetraborate decahydrate (Na₂BO₄), sodium chloride (NaCl), sodium nitrate (NaNO₃), sodium bicarbonate (NaHCO₃), sulphuric acid (H₂SO₄), glacial acetic acid (CH₃COOH) and sodium thiosulfate (Na₂S₂O₃) were purchased from Sigma-Aldrich (St. Louis, USA). Purified water was provided from JD Co. (Beijing, China). All the buffer solution was prepared using phosphate and borate to adjust the pH from 6.5 to 10.5.

2.2. Experimental procedures

The degradation experiments of CLQ were conducted in a 100 mL brown reagent bottle equipped with a magnetic stirring (500 r/min) at different temperatures (5, 10, 15, 20 and 25 °C). Reagents containing phosphate and borate were used to adjust the pH (7.3, 7.8, 8.2, 8.6, 9.0, 10.2) of the reaction solutions. The stock solutions of Fe(VI) (10 mM) were freshly prepared by dissolving solid K₂FeO₄ in pH 9.0 buffer (5 mM disodium hydrogen phosphate and 1 mM borate). The Fe(VI) stock solution was initially added to suspensions containing CLQ under rapid mixing. The initial concentration of Fe(VI) was controlled as 40, 90, 120, 150 and 180 μM while CLQ concentration was 10 μM. After reaction of 1.0, 3.0, 5.0, 7.0, 10.0, 15.0 and 20.0 min, 1 mL of the reaction solution was sampled and quenched with 0.1 mL sodium thiosulfate (3 mM) to detect residual CLQ concentrations. Meanwhile, 1 mL of the reaction solution was sampled to mix with 1 mL ABTS, 5 mL buffer solution (containing 0.6 M acetate and 0.2 M phosphate) and 19 mL purified water to detect the Fe(VI) concentrations by using the ABTS method at 415 nm [37]. Fig. S2 shows the adsorption capacity for CLQ by Fe(OH)₃ or Fe₂O₃, which formed by Fe(VI) decomposition. The result indicated Fe(OH)₃ or Fe₂O₃ formed by Fe(VI) decomposition had almost no adsorption capacity for CLQ.

The experiments of intermediate products identification were conducted in a 100 mL reaction suspensions of 10 μM CLQ and 150 μM Fe(VI) with pH 9.0 at room temperature (25 ± 1 °C). After 2 h, 60 mL samples were quenched with 1 mL sodium thiosulfate (10 mM) and stored at 4 °C for further analysis. All the experiments were carried out in triplicate. The average dates and the standard deviation were presented.

2.3. Analytical methods

The concentration of CLQ was determined by high-performance liquid chromatography (HPLC; Waters Corporation, USA) equipped with a Waters 2489 UV/Visible Detector and a XDB-C18 column (150 mm × 4.6 mm, 5 μm, Agilent Zorbax). The mobile phase for the measurement consisted of 0.1% (v/v) formic acid and acetonitrile at the ratio of 90:10 (v/v) with a flow rate of 1.0 mL/min. The detector temperature and wavelength were set at 40 °C and 340 nm, respectively. The limit of detection (LOD) for CLQ was 20 μg/L. The concentration of Fe(VI) was calculated using the ultraviolet–visible spectrophotometer (Beijing General Instruments Company Ltd, China) by sampling 1 mL solution to mix with 1 mL ABTS, 5 mL buffer solution (containing 0.6 M acetate and 0.2 M phosphate) and 19 mL deionized water to detect the Fe(VI) concentrations at 415 nm according to Eq. (1).

$${[Fe(VI)]}_{\mathit{sample}}=\frac{\mathrm{\Delta }{A}_l^{415}{V}_{\mathit{final}}}{\epsilon l{V}_{\mathit{sample}}}$$ (1)

The concentrations of intermediate products were extracted by liquid/liquid extraction with n-hexane (ultra resi-analysed grade) and then evaporated on a rotatory evaporator (temperature at 45 °C, rotor speed of 80 rpm) to a volume less than 2 mL. Then, the sample was identified by liquid chromatography-time-of-flight-mass spectrometry (LC-QTOF-MS/MS) equipped with a Waters cortecs C18 column (2.1 × 50 mm, 1.7 μm). The mobile phase consisted of 0.1% formic acid and acetonitrile (90:10, v/v) with 0.3 mL/min flow rate. The MS system was operated at ESI⁺ mode, capillary voltage 4 kV, gas temperature 150 °C, drying gas 15 L/min, sheath gas temperature 350 °C and sheath gas flow 12 L/min. The detection m/z range was from 100 to 500 and the scan time was 0.1 s.

2.4. Antibacterial activity assay

According to the previous study, the methods with E. coil were used for determining the biotoxicity of Fe(VI), CLQ and their oxidation products [38]. The experiment was carried out under the specific condition: 10 μM CLQ solution was reacted with Fe(VI) at the molar ratio of 18:1, the temperature was controlled as 25 °C. Four groups were set as follow: the control groups (adding 2 mL purified water into 20 mL of Luria-Bertani (LB) medium), the Fe(VI) decomposition groups (180 μM Fe(VI) dissolved in 100 mL purified water and decomposition for 12 h, then, 2 mL sample filtered through the hydrophilic acetate fiber membrane (0.45 μm) was added into 20 mL of LB medium), the CLQ groups (2 mL of sterile CLQ solution (10 μM) was added into 20 mL of LB medium), the experiment groups (Fe(VI):CLQ = 18:1, Fe(VI) reacted with CLQ for 12 h, 2 mL sample filtered through the hydrophilic acetate fiber membrane (0.45 μm) was added into 20 mL of LB medium). Then, E. coli in exponential phase was inoculated into the reactors. The optical density value of bacterial growth was determined by the absorbance of 600 nm using a UV–Visible spectrophotometry (T6 New century).

3. Results and discussion

3.1. Influence of Fe(VI) concentration on CLQ degradation

The fate for CLQ degradation was reacted with K₂FeO₄ in the various forms. To examine the role of the initial concentration of K₂FeO₄ on CLQ degradation, different molar ratios of [Fe(VI)]/[CLQ] varied between 4 and 18 were carried in the buffered solution at pH 9.0. As shown in Fig. 1 a, the degradation rate of CLQ increased with the increasing initial concentration of Fe(VI). When the molar ratios of [Fe(VI)]/[CLQ] were 4, 9, 12, 15 and 18, the degradation efficiencies of CLQ were 17.9%, 52.1%, 72.3%, 85.8% and 100% within 10 min, respectively. This phenomenon was attributed to the increased amount of Fe(VI) available to oxidize CLQ. Fig. 1b shows the final removal efficiencies of CLQ upon different molar ratios of [Fe(VI)]/[CLQ]. The final removal efficiencies of CLQ were 32.4% and 67.6% at the molar ratio of 4 and 9, respectively. At a molar ratio of 12:1 ([Fe(VI)]/[CLQ]), more than 90% of CLQ was eliminated. The results might be explained by considering simultaneous self-decomposition of Fe(VI) during the reaction (Fig. S3). The degradation of CLQ by Fe(VI) was slower than the simultaneous self-decomposition of Fe(VI), meaning that a high amount of Fe(VI) was needed to fully oxidize the CLQ. With the molar ratio greater than 12, the increase in removal became slower and almost complete removal of CLQ were observed at a molar ratio of 15 and 18. The half-lives at different molar ratios of [Fe(VI)]/[CLQ] were calculated using the solid line of Fig. 1b. In addition, Fig. S4 shows the extent of CLQ mineralization by Fe(VI) oxidation. The total organic matter (TOC) of the untreated water was 6.8 ± 0.1 mg/L. The TOC partial rapidly decreased after the Fe(VI) addition in the initial 1 h. Then, the TOC continued to slowly decrease with the prolonging of oxidation time. The higher the Fe(VI)/CLQ molar ratio, the more the TOC reduction. This result indicated only partial CLQ could be mineralized by Fe(VI) oxidation. Based on the data presented, the following pseudo-first-order rate (Eq. (2)) was proposed to verify further the effect of self-decomposition of Fe(VI) on the degradation of CLQ.

$$\mathit{ln}(\frac{{[CLQ]}_T}{{[CLQ]}_0})=-k_{\mathit{obs}}t$$ (2)

Fig. 1.

The effect of [Fe(VI)]/[CLQ] on the degradation efficiency of CLQ (a) and the effect of [Fe(VI)]/[CLQ] on the final removal efficiency of CLQ (b). (The ratio of [Fe(VI)]/[CLQ] = 4, 9, 12, 15, 18, initial CLQ concentration = 10 μM, pH = 9, T = 25 °C).

Plots of ln([CLQ]₀/[CLQ]_t) with t was established in Fig. S5. When the molar ratio of [Fe(VI)]/[CLQ] was more than 9, the linear correlation was higher than 0.946, which fitted the pseudo-first-order well. However, the linear correlation was unsatisfactory (R² = 0.882) when the molar ratio of [Fe(VI)]/[CLQ] was 4. This result was consistent with the above explanation about the self-decomposition of Fe(VI) (Fig. S3).

3.2. Influence of pH on CLQ degradation

Preliminary evaluation of CLQ degradation over a wide pH range (4–11) were conducted using the following buffers: citrate for pH 3.5, acetate for 5.5, phosphate and borate for 6.5–10.5. The instability of Fe(VI) under acidic conditions (pH 4.0–7.0) led to the rapid degradation of CLQ (data not shown). Therefore, alkaline conditions (pH = 7.0–11.0) for CLQ degradation were further investigated. The experiments carried out within the pH of 7.3, 7.8, 8.2, 8.6, 9.0 and 10.2. The molar ratio of ([Fe(VI)]/[CLQ]) was controlled as 15:1.

Previous studies have demonstrated that the reaction rate for the Fe(VI) oxidation with organic compound follows second-order reaction kinetics [24], [39]. Thereby, the reaction rate constant for the CLQ degradation by Fe(VI) can be expressed as

$$-\frac{d\left[CLQ\right]}{\mathit{dt}}=k_{\mathit{app}}[Fe(VI)][CLQ]$$ (3)

$$\mathit{ln}(\frac{[CLQ]}{[{[TCS]}_0]})=-k_{\mathit{app}}{\int }_0^t[Fe(VI)]dt$$ (4)

where k_app is the overall apparent reaction rate constant, [Fe(VI)] and [CLQ] are the concentration of Fe(VI) and CLQ, respectively. Fig. 2 a shows the values of k_app at different pH (7.3–10.2). Rates for the CLQ degradation by Fe(VI) decreased with an increase in pH, which was consistent with the reactions of antibiotics by Fe(VI) [33]. The rate constant was 7.7 × 10² M⁻¹s⁻¹ at pH 7.3 and decreased dramatically from pH 7.8 (k_app = 1.9 × 10² M⁻¹s⁻¹) to 10.9 (k_app = 18 M⁻¹s⁻¹). The values of k_app were pH-dependent, which could be attributed to the species of both Fe(VI) and CLQ. According to the Eqs. (5) to (9), in the pH range of 7.3–10.2, Fe(VI) and CLQ both have two acid–base species, which were HFeO₄ ^-, FeO₄ ^2- and CLQ^-, CLQ^2-, respectively.

$$H_{3}Fe{O}_4^{+}\to H_{2}Fe{O}_4+H^{+}p{K}_1=1.5(5)$$

$$H_{2}Fe{O}_4\to HFe{O}_4^{-}+H^{+}p{K}_2=3.5\left(6\right)$$

$$\mathit{HFe}{O}_4^{-}\to Fe{O}_4^{2-}+H^{+}p{K}_3=7.2\left(7\right)$$

$$\mathit{CLQ}\to CL{Q}^{-}+H^{+}p{K}_{a1}=8.1(8)$$

$$\mathit{CL}{Q}^{-}\to CL{Q}^{2-}+H^{+}p{K}_{a2}=10.2\left(9\right)$$

Fig. 2.

The effect of pH on the degradation efficiency of CLQ (a) and the model fitting result of the second-order reaction rate constant of ferrate with CLQ (b). (The ratio of [Fe(VI)]/[CLQ] = 15:1, initial CLQ concentration = 10 μM, pH = 7.3, 7.8, 8.2, 8.6, 9.0 and 10.2, T = 25 °C).

The CLQ degradation by Fe(VI) contributed to four possible reactions as follow.

$$\mathit{HFe}{O}_4^{-}+CLQ\to Fe\left(\mathit{III}\right)+Products{k}_1$$ (10)

$$\mathit{HFe}{O}_4^{-}+CL{Q}^{-}\to Fe\left(\mathit{III}\right)+Products{k}_2$$ (11)

$$\mathit{Fe}{O}_4^{2-}+CLQ\to Fe\left(\mathit{III}\right)+Products{k}_3$$ (12)

$$\mathit{Fe}{O}_4^{2-}+CL{Q}^{-}\to Fe\left(\mathit{III}\right)+Products{k}_4$$ (13)

Thus, the pH dependence of k_app can be calculated by Eq.(14).

$$-\frac{{d\left[Fe\left(VI\right)\right]}_{\mathit{tot}}}{\mathit{dt}}=k{\left[Fe\left(VI\right)\right]}_{\mathit{tot}}{\mathit{CLQ}}_{\mathit{tot}}=k_1\left[HFe{O}_4^{-}\right]\left[CLQ\right]+k_2\left[HFe{O}_4^{-}\right]\left[CL{Q}^{-}\right]+k_3\left[Fe{O}_4^{2-}\right]\left[CLQ\right]+k_4\left[Fe{O}_4^{2-}\right]\left[CL{Q}^{-}\right](14)$$

where k₁, k₂, k₃, and k₄ represent species-specific rate constants for reactions ((2), (3), (4)), respectively. The calculation of k_app can be simplified as (Eqs (15) and (16)):

$$k_{\mathit{app}}=\sum _{i=1,2j=1,2}{k}_{\mathit{ij}}{\alpha }_i{\beta }_j{\left[Fe\left(VI\right)\right]}_{\mathit{tot}}{\left[\left(CLQ\right)\right]}_{\mathit{tot}}\left(15\right)$$

$$k_{\mathit{app}}=k_{11}{\alpha }_1\left[HFe{O}_4^{-}\right]{\beta }_1\left[CLQ\right]+k_{12}{\alpha }_1\left[HFe{O}_4^{-}\right]{\beta }_2\left[CL{Q}^{-}\right]+k_{21}{\alpha }_2[Fe{O}_4^{2-}]{\beta }_1[CLQ]+k_{22}{\alpha }_2[Fe{O}_4^{2-}]{\beta }_2[CL{Q}^{-}](16)$$

where the α_i, β_j are the respective species distribution coefficient for Fe(VI) and CLQ while i, j each represents the corresponding two species of Fe(VI) and CLQ. The value of α_i, β_j can be calculated as follow.

$${\alpha }_1[HFe{O}_4^{-}]=[HFe{O}_4^{-}]/{[Fe(VI)]}_{\mathit{tot}}=[H^{+}]/([H^{+}]+{10}^{-7.23})(17)$$

$${\alpha }_2[Fe{O}_4^{2-}]=[Fe{O}_4^{2-}]/{\left[Fe\left(VI\right)\right]}_{\mathit{tot}}=[{10}^{-7.23}]/([H^{+}]+{10}^{-7.23})\left(18\right)$$

$${\beta }_1[CLQ]=[CLQ]/{[CLQ]}_{\mathit{tot}}=[H^{+}]/([H^{+}]+{10}^{-8.1})(19)$$

$${\beta }_1[CL{Q}^{-}]=[CL{Q}^{-}]/{[CLQ]}_{\mathit{tot}}=[{10}^{-8.1}]/([H^{+}]+{10}^{-8.1})(20)$$

Based on the least-squares nonlinear regressions of the experimental k_app data, the second-order rate constants of k₁, k₂, k₃, and k₄ were calculated as 1792, 82, 0 and 0 M⁻¹s⁻¹, respectively. The Fitting of the experimental data in Fig. 2b could describe very well with the experimental k_app (R² = 0.98). The aforementioned date indicated that: (1), no significant reaction between the HFeO₄ ⁻ and CLQ^- occurred and the reaction of HFeO₄ ⁻ controlled the overall rate of reaction; (2) the contribution of unprotoned ferrate and CLQ was not obvious and could be neglected. Therefore, the pH dependence of k_app for CLQ could be considered as the reactions of HFeO₄ ²⁻ with CLQ and CLQ^-. According to the different speciation of Fe(VI), the molar fraction of HFeO₄ ⁻ decreases and that of FeO₄ ²⁻ increases with the increasing pH [40]. Meanwhile, the protonated form of Fe(VI) has a larger spin density on the oxo ligands than the unprotonated form of Fe(VI), which increases the oxidation ability of protonated Fe(VI) [41]. In addition, the steadline in Fig. 2b shows that Eqs. (10), (11) were mainly contributed to the CLQ degradation. The protonated CLQ reaction with HFeO₄ ⁻ was faster than the unprotonated CLQ^- (k₂ > k₃).

3.3. Influence of temperature on CLQ degradation

The self-decomposition of Fe(VI) was related to the reaction temperature [42]. The effect of temperature on the CLQ degradation by Fe(VI) was varied from 5 °C to 25 °C at pH 9.0. As shown in Fig. 3 a. The rate of CLQ degradation increased with increasing temperature. According to the pseudo-first-order rates in Fig. S6, the k _obs of CLQ were increased from 0.034 M⁻¹min⁻¹ to 0.200 M⁻¹min⁻¹ while the temperature increased from 5 °C to 25 °C. There were two reasons contributed to this phenomenon. On the one hand, the increase of temperature would promote the Fe(VI) self-decomposing into the intermediate oxidants (Fe(VI)) and Fe(V)), which had more vital oxidation ability. In previous work, Wagner et al. found that Fe(VI) was much more stable at 0.5 °C than at 25 °C [42]. On the other hand, the increase in temperature increased the collision frequency between Fe(VI) and the target pollutant, which increased the elimination rate of CLQ [43]. Arrhenius equation was used to calculate the activation energy of CLQ degradation by Fe(VI) (Eq. (21)).

$$k_{\mathit{Obs}}=A{\mathit{exp}}^{{-}_{\mathit{RT}}^{E_a}}(21)$$

$${\mathit{ln}(k}_{\mathit{Obs}})=ln(A)-\frac{E_a}{\mathit{RT}}(22)$$

where R is the universal gas constant in 8.314 J/mol·K. Based on the experimental data, the rate constant for CLQ degradation can be estimated as follows:

$${\mathit{ln}(k}_{\mathit{Obs}})=-5137.4(\frac{1}{T})+21.972(23)$$

Fig. 3.

The effect of temperature (a), inorganic anions (b) and HA (c) on the degradation efficiency of CLQ. (d) The degradation of CLQ in natural waters (The ratio of [Fe(VI)]/[CLQ] = 15:1, initial CLQ concentration = 10 μM, pH = 9.0).

The ln k_obs and 1/T showed a satisfied correlation, which was higher than 0.97. The activation energy was calculated as 63.5 kJ/mol. Therefore, the higher temperature was prone to degradation the CLQ by Fe(VI).

3.4. CLQ degradation in natural water constituents

Various inorganic anions are ubiquitously existing in natural water, including Cl⁻, NO₃ ⁻ and HCO₃ ⁻, which can affect the environmental elimination of organic pollutants. In this study, the effects of three common inorganic anions (Cl⁻, NO₃ ⁻ and HCO₃ ⁻) on the CLQ degradation was evaluated by using the molar ratio of 15:1 ([Fe(VI)]:[CLQ]) at pH 9.0. The results were depicted in Fig. 3b.

Compared with the control experiment, the effect of Cl⁻ at the tested concentration (3 mM) had negative impact on CLQ removal at initial 5 min. This result indicated the fact that Cl⁻ concentrations might affect the dehalogenation during the oxidation of CLQ by Fe(VI), which could not facilitate the degradation of CLQ. Wu et al. also obtained the same conclusion when using Fe(VI) to degrade phenol in chloride ion water [44]. Another explanation might be that the pH value of the reaction solution was decreased with the presence of Cl⁻ which consumed more amount of Fe(VI) [45]. Besides, the effects of NO₃ ⁻ at the tested concentration (3 mM) slightly promoted the CLQ removal at initial 7 min. NO₃ ⁻ would act like to retard the decomposition of Fe(VI), and hence slightly favor Fe(VI) oxidation, which agreed well with the result of Schreyer et al. [9]. However, the effects of NO₃ ⁻ and Cl⁻ on the final degradation of CLQ by Fe(VI) were not noticeable. The removal of CLQ could reach 100% within 20 min. This was because of the low reactivity of Fe(VI) with NO₃ ⁻ and Cl⁻.Feng et al. [33] and Liu et al [46] found that the presence of NO₃ ⁻ and Cl⁻ could not affect the oxidation rate of flumequine and parathion, which was consistent with this result. Moreover, the addition of HCO₃ ⁻ significantly accelerated the oxidation of CLQ and Fe(VI). After reaction of 10 min, the removal rate of CLQ reached 100%. HCO₃ ⁻ raised the alkalinity and the ionic strength of the reaction system, enhancing the stability of Fe(VI) and improving the utilization efficiency of Fe(VI) [47].

Humic acid (HA) as an essential component of dissolved organic matters (DOM), is unavoidable in nature water. The effect of HA on removal of CLQ by Fe(VI) was shown in Fig. 3c. Compared with the control solution, the removal efficiency of CLQ by Fe(VI) was decreased with the addition of HA. When the solution contained 5 and 10 mg/L HA, the removal efficiencies were decreased to 73% and 63% within 7 min, respectively. HA decreased the oxidation of CLQ with Fe(VI), and the inhibitory effect increased with increasing HA concentrations. HA competed with CLQ to react with Fe(VI), which would consume Fe(VI) and inhibit the removal of CLQ by Fe(VI) in HA-containing water. This result was consistent with previous researches that HA interfered with removing organic chemicals (e.g., flumequine, benzophenone-3) and pharmaceuticals by Fe(VI) [39], [45].

In above section, HCO₃ ⁻ and HA would affect the degradation of CLQ by Fe(VI). In addition to HCO₃ ⁻ and HA, more background ions and dissolved organics existed in natural waters. It is critical to evaluate the feasibility of Fe(VI) oxidation technology for CLQ elimination in real natural waters. Fig. 3d shows the degradation of CLQ with Fe(VI) in tap water and river water. The degradation rate of CLQ decreased as follows: tap water > control water > river water. CLQ was completely removed from tap water and control water within 20 min. In the initial stage, the removal rate of CLQ was faster in tap water might due to the influence of the residual chlorine (Table S1) and coexisting anions (e.g., HCO₃ ⁻). In comparison, the removal efficiency of CLQ only surpassed 73% in river water, which might be attributed to the combined effects of coexisting anions and DOM (Table S1). The slight inhibition of the removal of CLQ in river water required adding more Fe(VI). Overall, Fe(VI) could be used as an efficient technique to degrade co-existing CLQ in natural waters.

3.5. Intermediate products and proposed degradation pathway

The complex transformation products formed during the reaction of CLQ and Fe(VI) were identified by using LC/ESI-QTOF-MS. The total ion current (TIC) chromatograms of the CLQ oxidation by Fe(VI) before and after 120 min are shown in Fig. 4 a. Meanwhile, frontier electron density theory and density functional theory (DFT) calculation were combined to determine the intermediate products (e.g., structure) and propose the degradation pathway. Fig. 4b shows the highest occupied molecular orbital (HOMO) of CLQ, and Table S2 shows the 2FED² _HOMO values of each atom of CLQ. The frontier orbital theory certified that electrophilic reactions were prone to occur at the atoms with higher 2FED² _HOMO values [48]. Therefore, according to the mass spectrums and the NIST MS software analysis, six intermediates were identified in Fig. 4c–h and Table S3.

Fig. 4.

HPLC/ESI-QTOF-MS chromatograms of the CLQ samples at different reaction time (a), the highest occupied molecular orbital of CLQ (b) and the extracted molecular ion mass/mass spectra of relevant oxidation products (c-h). (The ratio of [Fe(VI)]/[CLQ] = 15:1, initial CLQ concentration = 10 μM, pH = 9.0, T = 25 °C).

CLQ has a molecular ion peak of m/z = 320 ([M + H]⁺). Six intermediates were generated at m/z = 144, 179, 191, 247, 292 and 336. OP1 was identified with the m/z of 336 (M + H)⁺ in positive mode, which was 16 DA higher than CLQ (320 DA). The addition of 16 DA was corresponded to oxygen and was confirmed as (addition of one oxygen atom) mono N-oxide formation in CLQ, which was consistent with the research of Doddaga et al. [49] OP2 was identified with the m/z of 292 at 0.858 min, attributed to the loss of oxygen atom (–O, 16 DA) and loss of neutral ethylene (–C₂H₄) molecule from OP1. In addition, the peak at m/z 247 of OP3 was obtained the β-elimination of the aliphatic chain of the parent ion. The loss of neutral N,N-diethylamine hydroxide resulted in a terminal non-radicalic vinyl group, which protonated at the amine nitrogen. This result was also verified via quantum chemical calculations, where the orbital of HOMO isosurface was mostly distributed above the N,N-diethylamine (Fig. 4b). Besides OP1, OP2 and OP3, three products with m/z values of 191 Da (OP4), 179 Da (OP5) and 144 Da (OP6) were identified in Fig. 4f–h. The highest peak at 0.487 min was identified in the TIC chromatograms as an intermediate product (OP5, m/z = 179). Previous study has reported that OP5 could also be obtained by the reaction between CLQ and hydrogen peroxide [11]. The aromatic ring is prone to dealkylation, which leads to the breaking of the carbon–nitrogen bond [50]. The peak at 0.487 min was identified as OP4 (Fig. 4e), which had a mass of 12 Da higher than that of OP5. OP5 could be considered to be further oxidated by Fe(VI) to break carbon–nitrogen bond of the OP4. The m/z value of the compound arousing the peak at 0.779 min was OP6 (144 Da) (Fig. 4c), 35 Da lower than that of OP5. Based on the product OP6, the chlorine atom could be replaced by a hydroxyl group during the Fe(VI) oxidation of CLQ. The substitution of chlorine atoms was produced by the oxidation reactant of hydroxyl radicals generated by the self-decomposition of Fe (VI) in the buffer solution [37].

According to the above-identified intermediates and the theoretical calculations, a possible pathway of CLQ oxidation by Fe(VI) was proposed in Fig. 5 . The degradation of CLQ by Fe(VI) was mainly composed of aromatic ring dealkylation and chloride ion substitution. When Fe(VI) oxidizes organic matter, Fe(V), Fe(IV) and Fe(III) are gradually formed through single electron transfer. Fe(VI) initially attacked the N-diethylamine (electron rich moiety) of CLQ, in which N atom acted an electron pair donor to the high valent Fe(VI) via direct interaction with the ferrate oxygen atoms. The charge carried by N-diethylamine on CLQ was gradually transferred by Fe(VI) in the single electron to form mono N-oxide (OP1) and Fe(III). The reaction process was similar to the previous studies on sulfadiazine by Fe(VI) [51]. Besides, based on the higher 2FED² _HOMO value of N(2), electrophilic reactions were prone to occur at the atoms with higher 2FED² _HOMO values [48]. Subsequently, these compounds would transform into OP2 through loss of oxygen atom and ethylene (–C₂H₄) molecule. Due to protonation at the amine nitrogen of OP2, OP2 was further oxidized by Fe(VI) to form OP3. Then, the aromatic ring of OP3 was dealkylated to form OP4 and OP5. Finally, the chlorine atoms of CLQ could be replaced by a hydroxyl group by the oxidation of hydroxyl radicals produced by Fe(VI) self-decomposition to form OP6. Meanwhile, OP6 dealkylated the hydroxyl group to form OP7. Consistent with this work, CLQ was also observed dealkylation with the oxidation of hydrogen peroxide [49].

Fig. 5.

Proposed degradation pathway of CLQ with Fe(VI). (The ratio of [Fe(VI)]/[CLQ] = 15:1, initial CLQ concentration = 10 μM, pH = 9.0, T = 25 °C).

3.6. Antibacterial activity

Although CLQ could be effectively degraded by Fe(VI), its new intermediate products generated during the degradation process, such as OP6 and OP7, may possess potential biotoxicity. Hence, the CLQ degradation samples by Fe(VI) treatment were inoculated with E. coli to examine the antibacterial activity of the CLQ degradation products. Fig. 6 compares the effect of Fe(VI), CLQ and their oxidation products on the growth of E. coli. With the inoculation of E. coli for 24 h, the value of OD₆₀₀ were 1.355, 1.371 and 1.393 in the control group, Fe(VI) group and CLQ group, respectively. Compared with the control group, the growth of E. coli in CLQ group and Fe(VI) group were not obviously affected, indicating that the antimicrobial effect of CLQ (10 μM) on E. coli could be ignored. This might due to the fact that CLQ had no significant effect on DNA synthesis in E. coli. Meanwhile, Fe(VI) has no antibacterial effect on E .coli, which was consistent with the result of Yang et al [52]. However, after treatment by Fe(VI), the peak value of OD₆₀₀ in the CLQ degradation products was lower (1.249) than that of the control group (1.355), suggesting that the biotoxicity of CLQ degradation products towards E. coli appeared to be increased. Because the oxidation of CLQ by Fe(VI) further formed the lower molecular weight products containing benzene and pyridine, which exhibited higher toxicity than the CLQ group. EI-Sayed et al. reported that a series of pyridines as antimicrobial agents demonstrated varying antibacterial effects [53]. Therefore, this data further suggested that the CLQ degradation by ferrate would slightly increase the antimicrobial effect towards E. coli. In order to comprehensively evaluate the potential aquatic ecological risks of CLQ oxidation, the acute and chronic toxicity of CLQ and its intermediates was calculated by ECOSAR software (Fig. S8). According to the Globally Harmonized System of Classification and Labeling of Chemicals, CLQ and its intermediates have been classified into four categories (not harmful, harmful, toxic, and very toxic). In the acute toxicity, except OP-2, most of the intermediate products were less toxic to fish, daphnid and green algae than CLQ itself. Compared with the acute toxicity, the chronic toxicity of CLQ and intermediate products were increased. But the chronic toxicity of intermediate products was still less than CLQ itself. Thus, the potential toxicity risk of intermediates (e.g., acute toxicity and chronic toxicity) were weakened after the Fe(VI) oxidation.

Fig. 6.

Growth curve of E. coli in control group (diluted LB medium), CLQ group (10 μM), ferrate group (180 μM of ferrate dissolved in pure water, decomposed for 12 h) and ferrate treated CLQ group (10 μM of CLQ oxidized by 180 μM of ferrate for 12 h).

4. Conclusions

This study demonstrated that CLQ could be effectively oxidized by Fe(VI) within minutes in acidic to basic pH range. CLQ was degraded faster with the increasing of the Fe(VI) concentration, which was completely degraded by Fe(VI) with a stoichiometric ratio of 15:1 (Fe(VI):CLQ). The reaction was pH dependent and followed second-order kinetics. The species-specific rate constants determined for the reaction of HFeO₄ ⁻ with neutral CLQ and with anionic CLQ were 7.7 × 10² M⁻¹ s⁻¹ and 0.82 × 10² M⁻¹ s⁻¹, respectively. Besides, increasing the reaction temperature would increase the degradation rate of CLQ. Coexisting partial inorganic anions (e.g., HCO₃ ⁻) and HA could negatively influence the removal of CLQ by Fe(VI). Moreover, based on six oxidation products identified by LC-TOF-MS, a possible mechanism for the oxidation of CLQ by Fe(VI) was proposed, involving aromatic ring dealkylation and chloride ion substitution. Of note, toxicity evolution by luminescence tests and the antimicrobial effects toward E. coli showed that the oxidation of CLQ by Fe(VI) could slightly increase potential antibacterial activity.

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Appendix A. Supplementary data

The following are the Supplementary data to this article: