English

• Tetracycline antibiotics (TCs) are a class of broad-spectrum antibiotics produced by actinomycetes, mainly containing tetracycline (TC), oxytetracycline (OTC) and chlortetracycline (CTC) with a tetraphenyl basic skeleton [1]. TCs are widely used in the medical and animal husbandry industries due to their excellent antimicrobial properties and cost-effectiveness [2,3]. However, metabolism of TCs at low levels in organisms can result in their release into the aquatic environments as parent compounds (> 75%) and the concomitant high detection frequencies and concentrations in water. As one of the most widely used TCs, TC has been detected in wastewater, surface water and groundwater at concentrations of 2.2 mg/L, 0.1 µg/L and 4.5 µg/L, respectively [4]. Low concentrations of TC can lead to cytotoxicity and genetic resistance, threatening the ecosystem and human health [5,6]. In particular, the frequent detection of TC in medical wastewater may lead to bacterial resistance with potential harm to ecosystems [7]. Therefore, it is of significance to study the effective technologies for rapid TC abatement.

  Peracetic acid (PAA), generated via the reaction between acetic acid and H₂O₂ (Eq. S1 in Supporting information), has been widely used as the antibacterial agent and disinfectant after it was first synthesized in 1902 [8–10]. Compared to conventional chlorine disinfectants, PAA has comparable disinfection reaction activity and does not require dichlorination in the subsequent process [11,12]. The United States Environmental Protection Agency has approved the application of PAA disinfection for the treatment of sewage overflows in 1999 and for wastewater treatment in 2012 [12,13]. As a result, PAA is widely used in wastewater treatment, disinfection, medical, chemical and paper industries [8,14]. The excellent performance of PAA in disinfection application (especially in hospital wastewater disinfection) provides a good basis for its use in the chemical oxidation for the pollutant abatement.

  The degradation of pollutants is often achieved using advanced oxidation process that activate PAA to generate free radicals [13,15,16], ignoring the oxidative properties of PAA itself. Compared with conventional oxidants, PAA possesses higher standard oxidation potential (E^o =1.96 V) and therefore can selectively oxidize organic pollutants with specific structures in wastewater [12,17,18]. For instance, Zhang et al. and Du et al. found that PAA can attack the sulfur group of β-lactam and amino acids for its rapid oxidative degradation [19,20]. In addition, Kim and Huang summarized the ability of PAA to oxidize 123 organic compounds found that the rate constants of PAA with the compounds varied by about ten orders of magnitude [21], which is lower than O₃ [22], HOCl [23], but close to Fe(VI) [24]. Therefore, the reaction rate of PAA with organic contaminant greatly depends on the structure of the contaminant.

  Regarding that PAA is an electrophilic reagent and TC contains several electron-rich fractions, TC is expected to exhibit considerable reactivity toward PAA [21,25]. A previous study has demonstrated the reactivity between PAA and TC, yet the species-specific reactions for the reaction of PAA and TC are not clear, and the electron-donating and electron-accepting properties of the different species of TC and PAA needs to be elucidated. At the same time, the research method of combining experimental results with density functional theory (DFT) calculations, which can provide in-depth information with respect to the molecular level [26–29]. Is there a link between species-specific rate constants and the DFT calculations for the ability to gain or lose electrons of different TC and PAA species? Most importantly, considering that PAA is a commonly used oxidant in the disinfection of hospital wastewater and the frequent detection of TC in medical wastewater, oxidative of TC by PAA is practically feasible.

  Therefore, the objectives of this work are: (ⅰ) To evaluate the efficacy of oxidative abatement of TC by PAA for different pH conditions; (ⅱ) to construct the kinetic model and calculate the species-specific reaction rate constants for the reaction between PAA with TC; (ⅲ) to analyze the electron-donating and accepting property via the calculation of the highest occupied molecular orbital (HOMO), lowest unoccupied molecular orbital (LUMO), HOMO−LUMO gap, and electrostatic potential of TC and PAA in different forms based on DFT and (ⅳ) to identify the discrepancy of the reactive sites via the Fukui functions for different TC species.

  The detail information of chemicals was shown in the Text S1 in Supporting information.

  The batch experiments of TC abatement were performed in 250 mL glass vessel with 200 mL reaction solution and the rotary speed of 500 rpm. The experimental temperature was controlled at 25 ℃ using the water bath. The 10 mmol/L acetate buffer (3.5 < pH < 5), phosphate buffer (5 < pH < 8), borate buffer (pH > 8) or adding pre-calculated amounts of H₂SO₄ (3.0 ≤ pH < 3.5) were used to maintain the pH of the solution. TC was added in advance to the glass, followed by the addition of PAA concentration 10 times or higher than the initial TC concentration to initiate the reaction. The sample with 1 mL was withdrawn using the pipette gun at predetermined time intervals and subsequently added to the liquid phase vials containing 100 µL Na₂S₂O₄ (100 mmol/L) for TC concentration analysis. All experiments were repeated at least three time to minimize standard deviation.

  The methods of determination the PAA and H₂O₂ concentration in the PAA stock solution were shown in the Text S2 (Supporting information). The details of analytical methods and DFT calculation were provided in the Text S3 (Supporting information).

  Fig. S1 (Supporting information) presents TC abatement as a function of time by PAA under neutral conditions. TC does not undergo self-attenuation reaction, exhibiting stable properties. In the presence of 200 µmol/L PAA, rapid removal of TC (72.9%) is achieved within 60 min of reaction, and CH₃COOH and different concentrations of H₂O₂ in PAA solution exert negligible effect on the degradation of TC (Fig. S2 in Supporting information), indicating the redox reaction between the PAA and TC at pH 7.0. Comparison of kinetic data for the oxidation of TC by typical oxidants is shown in Table S1 (Supporting information). High TC abatement efficiencies can be achieved at low PAA concentrations, while the reaction between the PAA and TC is milder and reduces the generation of chlorine-containing by-products compared to chlorine oxidizers. To further explore the effect of pH, TC abatement is monitored under different pH conditions via controlling the ratio of PAA: TC. The specific experimental conditions are shown in Table S2 (Supporting information) (Fig. 1a). The difference in PAA concentration at different pH conditions is mainly for more accurate calculation of k_obs. The pH of the solution remains constant during the reaction (Fig. S3 in Supporting information). Subsequently, TC abatement by PAA is plotted by pseudo-first-order kinetics with high linear correlation coefficients (Fig. S4a in Supporting information) and then the pseudo-first-order rate constants (k_obs) of TC abatement are calculated (Fig. S4b in Supporting information). The TC abatement at different PAA concentrations follows pseudo-first-order kinetics (Eq. 1, Figs. S5a and b in Supporting information). The k_obs of TC abatement increases linearly with increasing PAA concentration. The ln k_obs vs. ln C_PAA plot presents a linear relationship with the slope close to 1, indicating that k_obs with PAA concentration is the first-order [25]. Therefore, the reactions of PAA and TC can be described in terms of second-order kinetics (Eq. 2). The apparent second-order rate constants (k_{2, app}) are determined in the pH range of 3.0−11.0 (Fig. 1b). As can be seen, TC abatement under alkaline conditions is more efficient than under acidic conditions. This is mainly due to the increased deprotonation of TC under alkaline conditions, where the electron density of deprotonated TC is higher than that of protonated TC, thus facilitating the attack of PAA [30]. However, as the pH continues to increase, the proportion of deprotonated PAA is the dominant species, leading to the decrease in its ability to TC oxidation, which is consistent with a previous report [25]. The decay of PAA with time at different pH is also determined in the presence and absence of TC (Figs. S6a and b in Supporting information). The PAA decay can be negligible (< 10%) under acidic to weakly basic conditions. However, at pH 10.0 and 11.0, the PAA decay exceeds 10%, which is mainly due to the self-decaying reaction of PAA under high pH conditions (Eqs. 3−5) [31,32].

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  Figure 1

  

  Figure 1.  (a) TC oxidation abatement by PAA under different pH. (b) The k_{2, app} of TC abatement under different pH. Experimental conditions: [PAA] = 200–3000 µmol/L, [TC] = 20 µmol/L, pH 3.0–11.0.

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  The reaction between PAA and TC can be expressed as a bimolecular reaction, which is closely related to their respective speciation (Figs. 2a and b) [33,34]. As illustrated in Fig. 1b, the k_{2, app} values of TC by PAA exhibits strong pH-dependency. k_{2, app} increases with increasing pH with the maximum around pH 8.5, and decreases rapidly when pH is further increased. The strong pH-dependency of k_{2, app} is due to the fact that PAA and TC have different species with different reactivities. Therefore, the species-specific kinetic model for the reaction of PAA and TC is constructed based on the equations listed in Table 1. TC has three pK_a-values (pK_a1 3.32, pK_a2 7.78 and pK_a3 9.58) corresponding to four different species, i.e., TTC⁺ (with one positive charge), TTC (neutral), TTC⁻ (with one negative charge) and TTC²⁻ (with two negative charge) (Fig. S7a in Supporting information). The amount of each component of TC at all pH conditions can be calculated by the Eqs. 6−10, which is displayed in Fig. 2a. Similarly, PAA with a pK_a value at pH 8.2 (Fig. S7b in Supporting information) possesses two species, protonated PAAH and deprotonated PAA⁻. The content of PAA at different pH can be expressed by Eqs. 11−13 and Fig. 2b. Then, substituting Eq. 6 and Eq. 11 into Eq. 2, and yields Eq. 14 and Eq. 15. k_{i, j} denotes the species-specific second-order rate constant between the i species of TC and the j species of PAA. The molar fractions of the different species of PAA and TC at different pH are obtained by ɑ_i × β_j, as illustrated in Fig. 2c.

  Figure 2

  

  Figure 2.  (a) TC speciation as a function of pH. (b) PAA speciation as a function of pH. (c) Effect of pH on the PAA-TC species (β_j × α_i) and kinetic modeling for the k_{2, app} of TC with PAA. (d) Species-specific reaction rate constants for PAA with TC. Experimental conditions: [PAA] = 200–3000 µmol/L, [TC] = 20 µmol/L, pH 3.0–11.0.

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  Table 1

  

  Table 1.  Kinetic modeling equations.

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  Based on the above analysis, the value of k_{i, j} is obtained by least squares regression of the experimental k_{2, app} with Eq. 15 via the user-defined nonlinear regression equation in the Origin software and the fitted curve is shown in Fig. 2c. The experimental k_{2, app} and modeled k_{2, app} can be well fitted and the species-specific rate constants can be correspondingly calculated, as shown in Fig. 2d. In the fitted data, the reaction rate constant between PAA⁻ and TTC⁺ is 0. At the same time, the mole fraction of PAA⁻ and TTC⁺ is extremely low (Fig. 2c), so the reaction between PAA⁻ and TTC⁺ can be negligible in the overall reaction. The reaction of protonated PAAH to each species of TC is significantly higher than that of the deprotonated PAA⁻. Especially, the species-specific reaction of PAAH with TTC²⁻ is significantly higher than that of the other species, which mainly contributes to the reaction between PAA and TC. For different TC species, the reactivity of PAA follows an order of TTC²⁻ > TTC⁻ > TTC. This is mainly because the higher deprotonation of TC means the higher electron density of ring structure. Interestingly, the reactivity between PAA⁻ and each species of TC is opposite to PAAH. This is mainly due to the electrostatic repulsion between the negative charge of deprotonated PAA⁻ and that of deprotonated TC, which results in a decrease in their reactivity.

  As mentioned earlier, the differences in the PAA and TC reaction rates of different species are due to differences in their ability to gain and lose electrons. In DFT calculations, the frontier orbital theory can reflect the ability of a molecule to gain or lose electrons ability of [35,36]. In a molecule, the electron on HOMO has the highest energy and is the least bound, which is the most active and easy to lose electrons. In comparison, LUMO has the lowest energy of all the unoccupied orbitals and is the easiest to accept electrons. These two orbitals determine the electron gain/loss and transfer ability of the molecule. Therefore, the frontier orbital theory is performed to further understand the electron gaining and losing abilities of different species of PAA and TC. Higher HOMO means the higher ability to lose electrons and tendency for oxidation and lower LUMO means the higher ability to gain electrons and tendency for reduction. In comparison, the HOMO–LUMO gap is the energy difference between HOMO and LUMO. A smaller HOMO–LUMO gap indicates lower energy required to transfer an electron from HOMO to LUMO and more reactive molecule, which promotes the occurrence of chemical reactions. The molecules with a large HOMO–LUMO gap tend to be more stable because the electron leaps are less likely due to the higher energy barrier [32]. The HOMO, LUMO and HOMO–LUMO gap of TTC⁺, TTC, TTC⁻, TTC²⁻, PAAH and PAA⁻ are calculated, as shown in Fig. 3. In terms of TC, the HOMO values of TTC⁺, TTC, TTC⁻, TTC²⁻ are –6.40 eV, –6.26 eV, –5.10 eV, –4.94 eV, respectively. The electron-losing capacity of TC increases with increasing deprotonation. For PAA, the LUMO values of PAAH and PAA^– are –0.24 eV and 0.60 eV, respectively. The electron-gaining capacity of PAA increases with increasing protonation. Meanwhile, the HOMO–LUMO gap of the four states of TC are 4.27 eV, 4.24 eV, 3.56 eV and 3.48 eV, representing that TC deprotonation can significantly facilitate the electrophilic attack by PAA, and in particular the HOMO–LUMO gap of TTC⁻ and TTC²⁻ are much smaller than TTC⁺ and TTC, which corresponds to the rate constants of the reaction.

  Figure 3

  

  Figure 3.  The HOMO, LUMO and HOMO–LUMO gap of (a) TC⁺; (b) TC; (c) TC^–; (d) TC^2–; (e) PAA; (f) PAA^–; (g) HOMO_TC–LUMO_PAA gap.

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  Previous studies have reported that the energy of HOMO–LUMO gap between two molecules can be used to evaluate their reactivity [37,38]. The smaller HOMO–LUMO gap indicates the larger polarizability, reflecting the excitation energy. In this study, the reactivity between different species of PAA and TC is evaluated to use the HOMO of different TC species minus the LUMO of different PAA species. As shown in Fig. 3g, the HOMO_TC–LUMO_PAA values are observed to follow an order of decreased by increasing deprotonation of TC and increased by protonation of PAA, which is in good agreement with the trend of species-specific rate constants (Fig. 2d). In fact, when the species-specific rate constant is calculated by experimental means, the electrostatic repulsion may exist when PAA and TC molecules in different states. However, in the defining HOMO_TC–LUMO_PAA, only the theoretical HOMO and LUMO of TC and PAA are taken into account, and the electrostatic repulsion present in PAA and TC is not taken into account when HOMO_TC–LUMO_PAA value is related to the actual species-specific reaction rate constants. Therefore, disregarding the effect of electrostatic interactions, the smaller HOMO_TC–LUMO_PAA value is, the bigger species-specific rate constant will be. Furthermore, the degree of deprotonation of TC significantly affects the HOMO_TC–LUMO_PAA gap. Similarly, the deprotonation of PAA leads to an increase of the HOMO_TC–LUMO_PAA gap with different species of TC. The smaller energy level of HOMO_TC–LUMO_PAA suggests the higher reactivity and the correspondingly larger reaction rate constants.

  The deprotonation of TC changes its electricity and affects its electron arrangement. Thus, the electrostatic potential (ESP) of TC is calculated, as shown in Fig. 4. Deprotonation of TC significantly affects its electron arrangement, potentially influencing the site at which its electrophilic reactions occur. In general, PAA is more inclined to electrophilic attack on the electron-rich sites of organic pollutants, and in particular, shows high reactivity towards the N-containing groups of phenolic compounds [21,25]. TC containing N groups and phenolic groups (Fig. S8 in Supporting information) possesses the four-rings structure and different degrees of protonation may alter their electron arrangement. The electrophilicity index ƒ ^– in the Fukui function can more accurately describe the electron-donating and accepting property on the atom. Therefore, the Fukui functions for TTC⁺, TTC, TTC⁻ and TTC²⁻ are calculated, as shown in Fig. S9 (Supporting information) and Tables S3−S6 (Supporting information), and the atomic numbers are shown in Fig. S10 (Supporting information). The easily attacked sites of four TC species are located at the carbon sites of the B and C rings, which is consistent with previous studies that TC are more susceptible to hydrogenation and hydroxylation reactions at the carbon sites [7,25]. Moreover, the sites susceptible to attack for different TC species is hardly changed, and thus the degree of deprotonation of TC does not affect its susceptibility to be attacked. Therefore, the protonation and deprotonation of TC does not change its oxidation products by PAA.

  Figure 4

  

  Figure 4.  The ESP of (a) TC⁺; (b) TC; (c) TC^–; (d) TC^2–.

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  In this work, the kinetics for the reaction between TC and PAA is explored, and the reactivity between different TC species and PAA is qualitatively evaluated by DFT calculations. The reactivity of PAA and TC is highly pH-dependent. Reactions of PAAH and TTC^2– are major contributors to reactions between PAA and TC. Electron-donating capacity of TC and electron-acquiring capacity of PAA exhibits a significant variation with increasing deprotonation level, which is proved by DFT calculations. The protonation of TC does not affect the reactive sites. Compared to conventional oxidizing agents, lower concentrations of PAA are required to achieve efficient TC degradation and no chlorine-containing by-products are produced. Hence, PAA oxidation exhibits significant advantages of the strong oxidizing ability of protonated PAA for TC abatement and the application prospects in wastewater treatment.

  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.

  Acknowledgments

  The authors wish to thank the National Natural Science Foundation of China (No. 52170088), Natural Science Foundation of Fujian Province (No. 2022J05064) and the Fundamental Research Funds for the Central Universities (No. ZQN-1118).

  Supplementary materials

  Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.cclet.2023.109436.

  
  
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• 

  Figure 1  (a) TC oxidation abatement by PAA under different pH. (b) The k_{2, app} of TC abatement under different pH. Experimental conditions: [PAA] = 200–3000 µmol/L, [TC] = 20 µmol/L, pH 3.0–11.0.

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  Figure 2  (a) TC speciation as a function of pH. (b) PAA speciation as a function of pH. (c) Effect of pH on the PAA-TC species (β_j × α_i) and kinetic modeling for the k_{2, app} of TC with PAA. (d) Species-specific reaction rate constants for PAA with TC. Experimental conditions: [PAA] = 200–3000 µmol/L, [TC] = 20 µmol/L, pH 3.0–11.0.

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  Figure 3  The HOMO, LUMO and HOMO–LUMO gap of (a) TC⁺; (b) TC; (c) TC^–; (d) TC^2–; (e) PAA; (f) PAA^–; (g) HOMO_TC–LUMO_PAA gap.

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  Figure 4  The ESP of (a) TC⁺; (b) TC; (c) TC^–; (d) TC^2–.

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  Table 1.  Kinetic modeling equations.

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