Degradation of Ketamine and Methamphetamine by the UV/H₂O₂ System: Kinetics, Mechanisms and Comparison

Abstract

The illegal use and low biodegradability of psychoactive substances has led to their introduction to the natural water environment, causing potential harm to ecosystems and human health. This paper compared the reaction kinetics and degradation mechanisms of ketamine (KET) and methamphetamine (METH) by UV/H₂O₂. Results indicated that the degradation of KET and METH using UV or H₂O₂ alone was negligible. UV/H₂O₂ had a strong synergizing effect, which could effectively remove 99% of KET and METH (100 μg/L) within 120 and 60 min, respectively. Their degradation was fully consistent with pseudo-first-order reaction kinetics (R² > 0.99). Based on competition kinetics, the rate constants of the hydroxyl radical with KET and METH were calculated to be 4.43 × 10⁹ and 7.91 × 10⁹ M⁻¹·s⁻¹, respectively. The apparent rate constants of KET and METH increased respectively from 0.001 to 0.027 and 0.049 min⁻¹ with the initial H₂O₂ dosage ranging from 0 to 1000 μM at pH 7. Their degradation was significantly inhibited by HCO₃⁻, Cl⁻, NO₃⁻ and humic acid, with Cl⁻ having relatively little effect on the degradation of KET. Ultraperformance liquid chromatography with tandem mass spectrometry was used to identify the reaction intermediates, based on which the possible degradation pathways were proposed. These promising results clearly demonstrated the potential of the UV/H₂O₂ process for the effective removal of KET and METH from contaminated wastewater.

1. Introduction

Illicit drugs are nonprescribed or psychostimulant substances which cannot be completely removed by conventional wastewater treatment, resulting in their widespread occurrence in aquatic environments [1,2]. Ketamine (KET) and methamphetamine (METH) were detected most frequently, with concentration levels up to 275 ng/L for KET and 239 ng/L for METH, in surface waters in China [3]. METH removal at most wastewater treatment plants was more than 80%, while the elimination of KET was less than 50% or even negative [4]. It was confirmed that chronic environmental concentrations of METH can lead to health issues in aquatic organisms [5]. Liao et al. [6] also reported that blood circulation and incubation time in medaka fish embryos could be significantly delayed at environmental concentration levels (0.004–40 μM) of KET and METH, which altered the swimming behavior of medaka fish larvae. Thus, there is an urgent need to explore new, efficient methods for eliminating these emerging contaminants in water.

Advanced oxidation processes (AOPs) have been employed to destroy illicit drugs due to their high efficiency and lower environmental impact [7,8]. The UV/H₂O₂ process is one of the AOPs and generates the strong, oxidizing hydroxyl radical (^•OH, E₀ = 2.72 V), which attacks the organic compounds with rate constants ranging from 10⁸ to 10¹⁰ M⁻¹ s⁻¹ [9]. Benzoylecgonine (BE), a metabolite of cocaine, was effectively removed by UV/H₂O₂ from different matrices [10]. The degradation of KET and METH was investigated using various AOPs, but no available report, so far, has addressed ^•OH assisted by UV/H₂O₂ treatment. After 3 min, 100 μg/L of METH that had been added to deionized water was completely eliminated by TiO₂ photocatalysis under UV_365nm irradiation [11]. Wei et al. [12] studied the synthesis of a novel sonocatalyst Er³⁺:YAlO₃/Nb₂O₅ and its application for METH degradation. Gu et al. [13] observed that complete removal of KET was achieved by UV/persulfate, and possible transformation pathways were proposed.

To the best of our knowledge, there is little information about the theoretical calculation of the reactivity of KET and METH by radical attack using the UV/H₂O₂ process. Water constituents in actual wastewater could affect the degradation efficacy; therefore, a comprehensive understanding of the degradation of KET and METH using the UV/H₂O₂ system is needed. The aim of this study was to investigate the degradation kinetics and mechanisms of KET and METH during the UV/H₂O₂ process. The influence of various parameters on KET and METH removal was evaluated, including initial H₂O₂ dosage, pH and water background components. The degradation products were analyzed by ultraperformance liquid chromatography with tandem mass spectrometry (UPLC-MS/MS), and possible transformation paths were proposed.

2. Materials and Methods

2.1. Materials

The KET and METH were obtained from Cerilliant Corporation (Round Rock, TX); detailed information is listed in Table 1. HPLC grade acetonitrile (ACN) and methanol (MeOH) were purchased from Fisher Scientific (Poole, UK). Formic acid (FA, ≥98%) and benzoic acid (BA) were purchased from Sigma-Aldrich (Bellefonte, USA). Analytical grade H₂O₂ (30%, v/v), NaHCO₃ (≥99.7%), NaCl (≥99.0%), NaNO₃ (≥99.5%), NaOH (≥99.5%), humic acid (HA) and H₂SO₄ (≥98%) were obtained from Sinopharm Chemical Reagent Co., Ltd. (Beijing, China). All reaction solutions were configured with Milli-Q water produced by an ultrapure water system (Millipore, MA, USA).

Table 1. Chemical structures and properties of ketamine and methamphetamine.

Compound	Chemical Formula	Structure	CAS Number	pKa	Log Kow
Ketamine	C13H16ClNO		6740-88-1	7.5	2.18
Methamphetamine	C10H15N		4846-07-5	9.9	2.07

2.2. Experimental Section

The experiments were operated in the quartz tubes (25 mm in diameter and 175 mm in length), which were placed in a photochemical reactor (Figure 1, XPA-7, Xujiang Machinery Factory, Nanjing, China). A low-pressure mercury lamp (11 W, emission at 254 nm, Philips Co., Zhuhai, China) was placed in the quartz sleeve. The UV lamp was preheated for 30 min to ensure irradiation stability. The UV fluence rate of 0.1 mW cm⁻² was determined using three different methods [14]. The newly configured KET/METH and H₂O₂ stock solutions were supplemented with appropriate volumes to achieve a 50 mL reaction solution, which was then stirred thoroughly at 300 rpm with electromagnetic stirrers. Upon UV irradiation, the reaction started at pH 7.0 and room temperature. Specific samples were immediately quenched using a catalase and passed through 0.22 μm nylon filter before further analysis.

Figure 1. The schematic diagram of the experiment setup: (1) low-pressure Hg UV lamp, (2) quartz tube, (3) cooling water, (4) photoreactor, (5) magnetic stirrer, (6) magnetic stirrer apparatus, (7) thermostat.

2.3. Analytical Methods

The concentrations of KET and METH were quantified by UPLC-MS/MS equipped with a Waters Acquity liquid chromatography system and an Xevo T_QS triple quadrupole mass spectrometer (Waters Co., Milford, MA, USA). The analytes were separated by a reverse phase column (Acquity UPLC BEH C18, 1.7 μm, 50 × 2.1 mm, Waters, MA, USA). The mobile phases A and B, with a flow rate of 450 μL min⁻¹, were 0.1% FA in Milli-Q water and ACN, respectively. Ten percent of phase B was kept for 0.5 min at the initial proportion, linearly increased to 45% at 1.8 min, then increased to 95% within 0.1 min, held for 1.0 min, reverted to 10% at 3.0 min and held for 1.5 min. The injection volume was 5 μL with the column temperature at 40 °C. The chromatograms were recorded in the positive ion multiple reaction monitoring (MRM) mode. Nitrogen was used as the desolvation and nebulizing gas. The capillary voltage was set at 0.5 kV, and the desolvation temperature was 400 °C. Optimized UPLC-MS/MS parameters are given in Table 2.

Table 2. Detailed ultraperformance liquid chromatography with tandem mass spectrometry (UPLC-MS/MS) parameters for ketamine and methamphetamine.

Compound	Parent Ion (m/z)	Retention Time (min)	Production (m/z)	Cone Voltage (V)	Collision Voltage (V)
Ketamine	238	1.31	125	16	24
			179	16	16
Methamphetamine	150	1.11	91	22	16
			119	22	10

3. Results and Discussion

3.1. Degradation Kinetics of KET and METH

Figure 2 shows the degradation of KET and METH under different treatment processes. UV or H₂O₂ alone exhibited negligible effects on their degradation, suggesting that treatment by UV or H₂O₂ alone was unable to destroy KET and METH. However, nearly complete removal of KET and METH was achieved within 120 and 60 min, respectively, when treated with the combination of UV/H₂O₂. Similar results were reported regarding ofloxacin degradation, which was drastically increased due to the large amount of hydroxyl radicals (^•OH) generated via the breakage of the H₂O₂ bond (Equation (1)) [15]. The degradation of KET and METH was consistent with the pseudo-first-order reaction kinetics. The apparent degradation rate constants (k_obs) of KET and METH by UV/H₂O₂ were 0.027 and 0.049 min⁻¹, respectively.

$$\mathrm{H}-\mathrm{O}-\mathrm{O}-\mathrm{H}+hv\to 2{}^{•}\mathrm{O}\mathrm{H}$$

(1)

Figure 2. Degradation kinetics of ketamine (KET) (a) and methamphetamine (METH) (b) by different treatments. Conditions: Initial concentrations of KET and METH = 100 μg/L, Initial concentration of hydrogen peroxide (H₂O₂)₀ = 500 μM, pH₀ = 7.0, Temperature (T) = 25 ± 1 °C.

3.2. Determination of Bimolecular Reaction Rate

The generation of ^•OH in the UV/H₂O₂ system was proved by the photoluminescence (PL) technique using a probe molecule with terephthalic acid, which tends to react with ^•OH to form 2-hydroxyterephthalic acid, a highly fluorescent product [16]. The PL intensity of 2-hydroxyterephtalic acid is proportional to the amount of ^•OH radicals produced in water [17]. Figure 3 shows the PL spectral changes in the 5 × 10⁻⁴ M terephthalic acid solution with a concentration of 2 × 10⁻³ M NaOH (excitation at 315 nm), as described by Yu et al. [17]. Similar fluorescence intensity was found in the reaction systems with initial concentrations of 100 and 1000 μM of H₂O₂, suggesting a constant concentration of ^•OH with the initial H₂O₂ dosage ranging from 100 to 1000 μM. The PL signal at 425 nm increased with the irradiation time, which was attributed to the reaction of terephthalic acid with ^•OH generated in the UV/H₂O₂ system.

Figure 3. Photoluminescence (PL) spectral changes observed in the UV/H₂O₂ system in a 5 × 10⁻⁴ M basic solution of terephthalic acid (excitation at 315 nm).

The bimolecular reaction rates of KET and METH reacting with ^•OH were determined through the competition experiments at pH 7 (phosphate buffer solution, 5 mM). BA was used as the reference compound, with which the constant reaction rate of ^•OH is known to be 5.9 × 10⁹ M⁻¹ s⁻¹ [18]. It is important to note that the degradation of KET, METH and BA using UV alone was negligible at less than 9%. Equations (2) and (3) describe the competing kinetics of KET and METH with ^•OH in the UV/H₂O₂ oxidation process, through which the bimolecular reaction rates of KET and METH reacting with ^•OH were 4.43 × 10⁹ and 7.91 × 10⁹ M⁻¹ s⁻¹, respectively (Figure 4).

$$\ln \frac{{\left(\mathrm{KET}\right)}_0}{{\left(\mathrm{KET}\right)}_{\mathrm{t}}}=\frac{k_{{•}_{\mathrm{O}\mathrm{H}-\mathrm{K}\mathrm{E}\mathrm{T}}}}{k_{{•}_{\mathrm{O}\mathrm{H}-\mathrm{B}\mathrm{A}}}}\ln \frac{{\left(\mathrm{BA}\right)}_0}{{\left(\mathrm{BA}\right)}_{\mathrm{t}}}$$

(2)

$$\ln \frac{{\left(\mathrm{METH}\right)}_0}{{\left(\mathrm{METH}\right)}_{\mathrm{t}}}=\frac{k_{{•}_{\mathrm{OH}-\mathrm{METH}}}}{k_{{•}_{\mathrm{O}\mathrm{H}-\mathrm{BA}}}}\ln \frac{{\left(\mathrm{BA}\right)}_0}{{\left(\mathrm{BA}\right)}_{\mathrm{t}}}$$

(3)

where (KET)₀, (METH)₀ and (BA)₀ are the initial concentrations (μmol/L) of target compounds. (KET)_t, (METH)_t and (BA)_t are the concentrations (μmol/L) at time t (min). k•_OH-KET, k•_OH-METH and k•_OH-BA are the bimolecular reaction rates of KET, METH and BA reacting with ^•OH, respectively.

Figure 4. (a) The reaction rate constant of KET with ^•OH. Conditions: (KET)₀ = (BA)₀ = 0.42 μM, (H₂O₂)₀ = 1 mM, pH = 7, T = 25 ± 1 °C. (b) The reaction rate constant of METH with ^•OH. Conditions: (METH)₀ = (BA)₀ = 0.67 μM, (H₂O₂)₀ = 1 mM, pH = 7, temperature = 25 ± 1 °C.

3.3. Effect of H₂O₂ Dosage

The KET and METH degradation under different initial H₂O₂ dosages were consistent with the pseudo-first-order reaction model (R² > 0.99, Figure 5). The k_obs of KET and METH increased dramatically from 0.001 min⁻¹ to 0.027 and 0.049 min⁻¹ with the initial H₂O₂ dosage ranging from 0 to 1000 μM. The reason for this phenomenon is that the production of ^•OH increased with the initial H₂O₂ dosage ranging from 0 to 1000 μM, thus accelerating the degradation rate of target compounds [19]. However, the k_obs of METH decreased slightly with the initial concentration of H₂O₂ increased to 2000 μM. A similar phenomenon was observed in a previous report that indicated that the degradation rates of cyclophosphamide and 5-fluorouracil were proportional to the H₂O₂ dosage and slightly decreased with excess H₂O₂ [20]. An excessive amount of H₂O₂ would cause the self-scavenging effect of ^•OH to form HO₂• and O₂⁻• (Equations (4) and (5)) [21], the low reactivity of which could reduce the degradation rate. Similar results were obtained concerning the degradation of ofloxacin [15] and chloramphenicol [22]. Moreover, large amounts of ^•OH were dimerized to H₂O₂, and the generated HO₂• and O₂⁻• subsequently participated in other reactions (Equations (6)–(9)) [23]. This negative effect was not observed in this study, probably because the maximum H₂O₂ dosage (2000 μM) was not high enough to inhibit the KET degradation.

$${\mathrm{H}}_2{\mathrm{O}}_2{+}^{•}\mathrm{OH}\to {\mathrm{HO}}_2^{•}+{\mathrm{H}}_2\mathrm{O}$$

(4)

$${\mathrm{H}}_2{\mathrm{O}}_2{+}^{•}\mathrm{OH}\to {\mathrm{O}}_2^{-•}+{\mathrm{H}}^{+}+{\mathrm{H}}_2\mathrm{O}$$

(5)

$${}^{•}\mathrm{OH}{+}^{•}\mathrm{OH}\to {\mathrm{H}}_2{\mathrm{O}}_2$$

(6)

$${\mathrm{HO}}_2^{•}+{\mathrm{H}}_2{\mathrm{O}}_2\to {}^{•}\mathrm{O}\mathrm{H}+{\mathrm{H}}_2\mathrm{O}+{\mathrm{O}}_2$$

(7)

$${\mathrm{HO}}_2^{•}{+}^{•}\mathrm{OH}\to {\mathrm{H}}_2\mathrm{O}+{\mathrm{O}}_2$$

(8)

$${\mathrm{O}}_2^{-•}+{\mathrm{H}}_2{\mathrm{O}}_2{\to }^{•}\mathrm{OH}+{\mathrm{OH}}^{-}+{\mathrm{O}}_2$$

(9)

Figure 5. Effect of H₂O₂ dosage on KET (a) and METH (b) degradation in the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 0–2000 μM, pH₀ = 7.0, T = 25 ± 1 °C.

3.4. Effect of Initial pH

Figure 6 illustrates the KET and METH destruction at different initial pHs, which were adjusted with an H₂SO₄ or NaOH solution (0.1 M). No buffer was used due to its inhibiting effect on the decomposition of organics [24]. The KET and METH degradation at different initial pHs followed the pseudo-first-order reaction model well. The k_obs of KET and METH reached the highest levels in a neutral environment at 0.027 and 0.085 min⁻¹, respectively. Due to the greater stability of H₂O₂ at pH 5 and 7, the degradation rates of KET and METH under acidic and neutral conditions were obviously better than those under alkaline conditions. Under alkaline conditions, ^•OH could be quenched by the HO₂⁻ produced by H₂O₂ dissociation, thus reducing the yield of ^•OH in the system.

Figure 6. Effects of different initial pHs on the degradation of KET (a) and METH (b) in the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 3–11, T = 25 ± 1 °C.

3.5. Effect of Water Background Components on Degradation Efficiency of Target Compounds

There are many different substrates in natural water, including different kinds of anions, cations and organic matter. These ions could react with free radicals in advanced oxidation processes, thus inhibiting or promoting the reaction and affecting the overall oxidation effect. Therefore, it is of great significance to study the influence of different ion types and contents on the practical application of advanced oxidation technology.

3.5.1. Effect of HCO₃⁻

The decomposition of KET and METH was significantly inhibited with the addition of HCO₃⁻ at different initial dosages in the UV/H₂O₂ oxidation process (Figure 7). When the initial dosage of HCO₃⁻ ranged from 0 to 10 mM, the reaction rate of KET and METH decreased from 0.027 and 0.049 min⁻¹ to 0.008 and 0.011 min⁻¹, respectively. The reason for this experimental phenomenon was that HCO₃⁻ was the quenching agent for ^•OH which was also consumed by the competing reaction of ionized CO₃²⁻ (Equations (10)–(13)). Therefore, the inhibitory effect of KET and METH degradation was more obvious with the increase of the HCO₃⁻ concentration.

$${\mathrm{CO}}_3^{2-}+{}^{•}\mathrm{O}\mathrm{H}\to {\mathrm{CO}}_3^{-•}+{\mathrm{OH}}^{-}$$

(10)

$${\mathrm{HCO}}_3^{-}+{}^{•}\mathrm{O}\mathrm{H}\to {\mathrm{HCO}}_3^{•}+{\mathrm{OH}}^{-}$$

(11)

$${\mathrm{HCO}}_3^{•}\to {\mathrm{CO}}_3^{•}+{\mathrm{H}}^{+}$$

(12)

$${\mathrm{CO}}_3^{-•}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{HCO}}_3^{-}+{\mathrm{HO}}_2^{•}$$

(13)

Figure 7. Effect of HCO₃⁻ on KET (a) and METH (b) degradation in UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 7.0, T = 25 ± 1 °C.

3.5.2. Effect of Cl⁻

With the initial concentration of Cl⁻ ranging from 0 to 10 mM, the destruction of KET was dramatically inhibited with the rate constant of KET decreased from 0.027 to 0.018 min⁻¹ (Figure 8), which could be due to the elimination of ^•OH by Cl⁻ according to Equations (14)–(16) [25]. The degradation reaction rate changed slightly as more Cl⁻ was added. However, the METH degradation was less affected by Cl⁻, with the reaction rate remaining basically unchanged (0.0446–0.0485 min⁻¹).

Figure 8. Effect of Cl⁻ on KET (a) and METH (b) degradation in the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 7.0, T = 25 ± 1 °C.

3.5.3. Effect of NO₃⁻

$${}^{•}\mathrm{O}\mathrm{H}+{\mathrm{Cl}}^{-}\to {\mathrm{Cl}}^{•}+{\mathrm{OH}}^{-}$$

(14)

$${\mathrm{Cl}}^{•}+{\mathrm{Cl}}^{-}\to {\mathrm{Cl}}_2^{-•}$$

(15)

$${\mathrm{Cl}}^{•}+{\mathrm{Cl}}^{•}\to {\mathrm{Cl}}_2$$

(16)

The influence of NO₃⁻ on the decomposition of KET and METH is illustrated in Figure 9. With the initial concentration of NO₃⁻ ranging from 0 to 10 mM, the degradation of both target compounds was obviously inhibited. The reaction rate of KET and METH decreased from 0.027 and 0.049 min⁻¹ to 0.007 and 0.012 min⁻¹, respectively. The above experimental phenomena were attributed to the following: First, a large amount of ^•OH could be produced from NO₃⁻ under UV irradiation (Equations (17)–(18)), which is an important source of ^•OH in natural water [26]. Second, as a photosensitizer, NO₃⁻ has a strong absorption in the ultraviolet range, which results in the formation of an internal filter that prevents the effective light transmittance and leads to the decline of ^•OH production in the UV/H₂O₂ system [27]. The latter was found to be dominant after the degradation effect of the reaction was analyzed.

$${\mathrm{NO}}_3^{-}+hv\to {\mathrm{NO}}_2^{-•}+{\mathrm{O}}^{-•}$$

(17)

$${\mathrm{O}}^{-•}+{\mathrm{H}}_2\mathrm{O}\to {}^{•}\mathrm{O}\mathrm{H}+{\mathrm{OH}}^{-}$$

(18)

Figure 9. Effect of NO₃⁻ on KET (a) and METH (b) degradation in the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 7.0, T = 25 ± 1 °C.

3.5.4. Effect of HA

Due to its complex structure, HA may have uncontrollable effects on the destruction of target compounds. As illustrated in Figure 10, KET and METH degradation was dramatically inhibited once HA was added with different dosages in the UV/H₂O₂ system. As more HA (0–0.1 mM) was added, the reaction rate of KET and METH declined from 0.027 and 0.049 min⁻¹ to 0.001 and 0.008 min⁻¹, respectively, while the degradation reaction rate changed slightly with the continued addition of the HA. UV irradiation was absorbed by HA, creating an inner filter (Figure 11) and significantly inhibiting the UV transmittance for UV photons, thus limiting the generation of ^•OH in the UV/H₂O₂ process [28]. Moreover, the degradation of target compounds can be inhibited by the competing reaction of HA with the active radicals [29].

Figure 10. Effect of HA on KET (a) and METH (b) degradation in the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 7.0, T = 25 ± 1 °C.

Figure 11. The ultraviolet–visible spectroscopy of reaction solutions at different concentrations of HA.

3.6. Degradation Products and Mechanism

Degradation intermediates and products of METH produced in the UV/H₂O₂ oxidation process were determined by using UPLC/MS/MS under full scans and product ion scans. During the whole METH degradation process, the mass spectra were compared to identify the intermediates. The structure of the transformation products was analyzed with the specific molecular ions and fragmentation patterns rather than direct comparison with corresponding standards. Figure 12 illustrates the mass spectra and possible structures of the degradation intermediates, based on which the possible transformation pathways of METH during UV/H₂O₂ are shown in Figure 13. The proposed degradation mechanisms of METH degradation involved in the UV/H₂O₂ system include hydrogenation, hydroxylation and electrophilic substitution.

Figure 12. Mass spectra of the intermediate products of METH in the UV/PS system.

Figure 13. Tentative transformation of METH pathways in the UV/H₂O₂ system.

With the molecular weight of 149, intermediate product 2 (P2, m/z = 150) was formed as a result of hydrogenation of METH. P1 (m/z = 91) with a stable structure was generated from the fracture of the C-C bond of the branched chain. Intermediates P3 (m/z = 110) and P4 (m/z = 73) were formed by electrophilic substitution of hydroxyl. METH was hydroxylated to form ephedrine (m/z = 165), of which the C-C bond of branched chain was fractured to form intermediate product P5 (m/z = 57). The hydroxylation of ephedrine induced the formation of intermediate P6 (m/z = 181) which was then achieved to form intermediate P7 (m/z = 89) after further hydroxylation. The mineralization of KET and METH was characterized by removal of total organic carbon (TOC), which achieved 41% and 57% within 60 min under UV/H₂O₂ treatment (Figure 14). The intermediate products were further degraded as the reaction continued.

Figure 14. The mineralization of KET and METH during the UV/H₂O₂ system. Conditions: (KET)₀ = (METH)₀ = 100 μg/L, (H₂O₂)₀ = 500 μM, pH₀ = 7, T= 25 ± 1 °C.

4. Conclusions

The degradation kinetics and mechanisms of KET and METH using the UV/H₂O₂ process were investigated in this study. Their degradation in UV photolysis or H₂O₂ oxidation alone was negligible. However, 99% of KET and METH (100 μg/L) were effectively eliminated by the combination of UV and H₂O₂ within 120 and 60 min, respectively. According to the competition kinetics, the rate constants of ^•OH with KET and METH at pH 7 were calculated to be 4.43 × 10⁹ and 7.91 × 10⁹ M⁻¹·s⁻¹, respectively. The apparent rate constants of KET and METH reached the highest levels in a neutral environment. The degradation of KET and METH was significantly inhibited by HCO₃⁻, Cl⁻, NO₃⁻ and HA; however, Cl⁻ had little influence on the METH degradation. Seven reaction intermediates of METH in the UV/H₂O₂ system were identified by UPLC-MS/MS. Possible transformation mechanisms involved in the KET and METH degradation by UV/H₂O₂ oxidation system included hydrogenation, hydroxylation and electrophilic substitution. Results demonstrated that UV/H₂O₂ was an effective technique to remove KET and METH, providing a promising application for the decomposition of trace organic pollutants in natural water.