- Email: [email protected]
Removal of methamphetamine by UV-activated persulfate: Kinetics and mechanisms
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
Contents lists available at ScienceDirect
Journal of Photochemistry & Photobiology A: Chemistry journal homepage: www.elsevier.com/locate/jphotochem
Removal of methamphetamine by UV-activated persulfate: Kinetics and mechanisms
T
Deming Gua,b, Changsheng Guoa, Jiapei Lva, Song Houa, Yan Zhanga, Qiyan Fengb, Yuan Zhanga, ⁎ Jian Xua, a b
State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing 100012, China School of Environment Science and Spatial Informatics, China University of Mining and Technology, Xuzhou 221116, China
ARTICLE INFO
ABSTRACT
Keywords: Methamphetamine Persulfate Sulfate radical Hydroxyl radical Advanced oxidation processes (AOP)
In this study, removal of methamphetamine (METH) was investigated by UV activated persulfate (PS), and the influence of key factors was evaluated. Results suggested that METH degradation followed pseudo-first order reaction kinetics. The combination of UV and persulfate (UV/PS) could completely degrade 100 μg/L of METH in 30 min with a PS dosage of 200 μM at pH 7. Both hydroxyl radical (%OH) and sulfate radical (SO4−%) were confirmed to contribute to the degradation of METH. The bimolecular reaction rate constants of METH with %OH and SO4−% were 7.91 × 109 and 3.29 × 109 M−1 s−1, respectively. The degradation rate constant of METH was proportional to the PS dosage (0–800 μM) and was high at neutral pH condition. The presence of inorganic anions significantly reduced METH degradation to different degrees, with the inhibitory effect order of Cl− > NO3− > HCO3−. The degradation efficiency of METH was suppressed by the presence of humic acid due to the effect of UV absorption and free radical quenching. The degradation intermediates and products were identified by UPLC-MS/MS and possible transformation pathways were proposed. Results suggested that the combination of UV/PS is a promising treatment technique for the removal of METH in the water environment.
1. Introduction Methamphetamine (METH) is a commonly consumed drug of abuse. According to World Drug Report 2018, METH contributes a large proportion to the synthetic drugs in the world illicit drugs market [1]. After being consumed, METH cannot be completely metabolized in the body, and the parent METH and its metabolites could be excreted into the environment via urine and feces. In the conventional wastewater treatment plants, the complete elimination of METH can rarely be achieved [2], leading to its frequent detection in different water bodies around the world [3,4]. For instance, a recent survey covering 49 lakes and 4 major rivers across China revealed that METH and ketamine were found with high levels and detection frequency [5]. In the urban rivers in Beijing, our previous study showed that METH was one of the most predominant drugs with detection frequency ranging from 65% to 100% [6]. Although METH was observed at trace concentrations in the natural aquatic ecosystem, it may pose negative effect to both humans and other organisms via long-term exposure [7,8]. Many measures have been attempted to treat METH from aqueous solution. Rodayan et al. [9] reported that ozonation was capable of
⁎
eliminating drugs of abuse including METH from wastewater, with the generation of some persistent intermediates. METH can be efficiently removed by TiO2 photocatalysis under UV365 irradiation, with a complete elimination of METH in 3 h with the dosage of TiO2 at 0.1 g/L [10]. Fenton and Fenton-like reactions were also effective to degrade psychoactive pharmaceuticals and illicit drugs to below the detection limit in the wastewater [11]. Activated persulfate oxidation is also one of the advanced oxidation processes (AOPs) that has drawn great interest because of the distinctive advantage of sulfate radical (SO4−%) with high stability, selectivity, redox potential (2.5–3.1 V) and applicability in wide pH ranges [12]. SO4−% could be produced from persulfate under the activated conditions of heat, light, transition metal, alkaline environment, activated carbon, electrochemical, complexing agent and other new types [13]. Of all the treatment techniques, UV/PS process is deemed as a costly, sustainable and environment-friendly method, which can generate SO4−% with a quantum yield of 0.7 mol E/s (Eq. (1)) [14], and other subsequent primary chemical reactions (Eqs. (2)–(5) [15].
S2 O82
Corresponding author. E-mail address: [email protected] (J. Xu).
https://doi.org/10.1016/j.jphotochem.2019.05.009 Received 1 March 2019; Received in revised form 20 April 2019; Accepted 5 May 2019 Available online 06 May 2019 1010-6030/ © 2019 Elsevier B.V. All rights reserved.
UV
2SO4 •
= 0.7 molE/s
(1)
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
SO4 • + OH
SO24 +•OH k = 6.5 × 107M 1s
SO4 • + H2 O
HSO24 +•OH k = 8.3M 1s
• OH+S2 O82
S2 O•8 +OH
• OH+SO•4
1
(3)
1
k = 1.4 × 107M 1s
HSO5 k = 1 × 1010M 1s
(2)
1
1
(4) (5)
The combination of UV and PS has been confirmed to destruct a range of contaminants, such as indomethacin [16], 1H-benzotriazole [17], microcystis aeruginosa [18] and perfluorooctanoic acid [19]. To our best knowledge, no information is available on the elimination of METH by UV/PS. The aim of the present study is to (1) investigate the degradation and mechanism of METH in the UV/PS system under various experimental condition, (2) determine the bimolecular reaction rate constants of METH with %OH and SO4−% through competitive kinetics, and (3) identify the oxidation intermediates. 2. Materials and methods Fig. 1. Removal of METH under different oxidation treatment. [METH]0 = 100 μg/L, initial pH = 7.0, PS dosage was 200 μM, temperature = 25 ± 1 °C.
2.1. Materials METH was obtained from Cerilliant Corporation (Round Rock, TX, USA), with detailed information listed in supplementary materials Table S1. HPLC grade acetonitrile and methanol (MeOH) were obtained from Fisher Scientific (Poole, UK). Sodium persulfate (PS, 98.0%), formic acid (FA, ≥98%), tert-butyl-alcohol (TBA) and benzoic acid (BA) were obtained from Sigma-Aldrich (Bellefonte, USA). Analytical grade hydrogen peroxide (H2O2, 30%, v/v), sodium bicarbonate (NaHCO3, 99.7%), sodium chloride (NaCl, ≥99.0%), sodium nitrate (NaNO3, ≥99.5%), humic acid (HA), Sulfuric acid (H2SO4, ≥98%) and sodium hydroxide (NaOH, ≥99.5%) were obtained from Sinopharm Chemical Reagent Co., Ltd. (Beijing, China) and used without further purification. All solutions were prepared using Milli-Q water produced by an ultrapure water system (Millipore, MA, USA).
USA). The mobile phase comprised of 0.1% FA in Milli-Q water (phase A) and acetonitrile (phase B), and the flow rate was kept at 450 μL min−1. Initial proportion of phase B (10%) was held for 0.5 min, and linearly increased to 45% within 1.8 min, then increased to 95% at 1.9 min, held for 1.0 min, reverted to the initial condition (10%) within 0.1 min, and kept for 1.5 min. 5 μL of analytes were injected, with the column temperature at 40 °C. The positive ion multiple reaction monitoring (MRM) mode was adopted to record chromatograms. The capillary voltage was set at 0.5 kV. Desolvation temperature was 400 °C and source temperature was 150 °C. Nitrogen was used as the desolvation and nebulizing gas. Detailed UPLC-MS/MS parameters were given in Table S2. A UV–vis spectrophotometer was used to measure the concentration of residual PS [22].
2.2. Experimental section
3. Results and discussion
The experiments were conducted in a photo-reactor (Fig. S1) with an effective volume of 50 mL (XPA-7, Xujiang Machinery Factory, Nanjing, China). A low-pressure mercury lamp (11 W, Philips Co., China) was placed in the quartz sleeve. A filter has been used to allow only 254 nm wavelength transmission. To keep the UV irradiation stable, the lamp was warmed up before experiment. The UV fluence rate of 0.1 mW cm−2 was determined by three different methods [20]. The quantum yield of METH at 254 nm was determined to be 0.013 Einstein-1 [21]. Appropriate volumes of freshly prepared METH and PS stock solutions were added to achieve 50 mL reaction solution. The homogeneous mixtures were then stirred thoroughly with electromagnetic stirrers (300 rpm). The reaction solution was moved under UV irradiation to initiate the reaction at pH 7.0 and room temperature. Samples were taken at specific time intervals and quenched immediately using the same volume of methanol and passed through 0.22 μm nylon filter prior to further analysis. The initial pH values were adjusted with H2SO4 or NaOH solution (0.1 M). MeOH or TBA was added respectively as scavengers for %OH or SO4-% to investigate the reactive species involved in the UV/PS reaction system.
3.1. Removal of METH by UV, PS and UV/PS The removal of METH by UV, PS and UV/PS oxidation was illustrated in Fig. 1. METH degradation well fit the pseudo-first order model (R2 > 0.95, Eq. (6)). Negligible removal of METH was observed by direct UV irradiation or PS oxidation alone, with the same degradation rate constant of 0.005 min−1. In contrast, METH was completely removed within 30 min under UV/PS treatment. The generation of %OH and SO4-% in the UV/PS system was likely responsible for the enhanced elimination of METH. The pseudo-first order rate constant of METH degradation by UV/PS was calculated to be 0.193 min−1. Slight changes (less than 15%) were observed for the concentration of remaining PS during PS oxidation alone or UV/PS system. The mineralization of METH was characterized by removal of TOC, which achieved 44% within 30 min under UV/PS treatment (Fig. S2).
In
[METH] [METH]0
=
kobs t
(6)
where [METH] is the concentration of METH at time t (min), [METH]0 is the initial concentration, and kobs is the pseudo-first order rate constant (min−1).
2.3. Analytical methods The concentration of METH was quantified by an ultra-performance liquid chromatography tandem mass spectrometry (UPLC-MS/MS) consisting of a Waters ACQUITY liquid chromatography and an Xevo T_QS triple quadrupole mass spectrometer (ESI-MS/MS, 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,
3.2. Determination of bimolecular reaction rate constants Radical scavengers were used to identify the predominant radical species. The results indicated METH degradation was attributed to both 33
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
OH and SO4−% oxidation (detailed description in Text S1). The bimolecular reaction rate constants of METH with %OH and SO4−% were determined by competition kinetics at neutral pH. Benzoic acid (BA) was used as the probe chemical, and the constant reaction rates of %OH and SO4−% with BA were 5.9 × 109 M-1 s-1 and 1.2 × 109 M-1 s-1, respectively [23]. Less than 10% of METH and BA were degraded under UV irradiation alone, thus direct photolysis was negligible. The METH degradation kinetics in UV/PS system could be calculated with the following Eq. (7):
%
d[METH] = kSO4 • dt
METH [METH][SO4
•
] + k•OH
METH [METH][•OH]
(7)
TBA (50 mM) was added into the solution for scavenging hydroxyl radical in the UV/PS system. METH and BA were degraded only by SO4−%. The bimolecular reaction rate constant of METH with SO4−% could be obtained by Eq. (8):
kSO4 •
METH
= kSO4 •
BA
In([METH]t /[METH]0 ) In([BA]t /[BA]0 )
(8)
Fig. 3. Effect of initial pH on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
As above, the bimolecular reaction rate constant of METH with OH at pH 7 in UV-254 nm/H2O2 system can be obtained with the following Eq. (9): %
k•OH
METH
= k•OH
In([METH]t /[METH]0 ) BA In([BA]t /[BA]0 )
SO4 • + S2 O28
(9)
SO4 • + SO4 •
The plot of ln([METH]0/[METH]t)/ln([BA]0/[BA]t) yielded a straight line with a slope of 1.34 and 2.74, as presented in Fig. 2. The bimolecular reaction rate constants of METH with %OH and SO4−% at pH 7 were 7.91 × 109 and 3.29 × 109 M-1 s-1, respectively. It is reported that the reaction rate constants of organic compounds by %OH varied from 106 to 1010 M-1 s-1 [24], suggesting the rapid reaction with %OH.
S2 O8 • + SO24
S2 O82
k = 1.2 × 106M 1s
k = 4 × 10 8M 1s
1
1
(10) (11)
3.4. Effect of solution pH Solution pH may affect the degradation rates because of the difference of electronic distribution in molecules. As shown in Fig. 3, METH degradation in UV/PS system was dependent on pH, with the kobs values following the order of pH 7.0 (0.193 min−1) > pH 3.0 (0.125 min−1) > pH 5.0 (0.089 min−1) > pH 11.0 (0.032 min−1) > pH 9.0 (0.021 min−1). Overall, the reaction rates under acidic and natural conditions were higher than that under base conditions, which was likely attributed to the low activation energy of SO4-% under neutral pH condition [29]. Moreover, SO4-% was able to react with OH- to produce lower redox potential of %OH under alkaline conditions (Eq. (12) [28], which might lessen the contribution of SO4-% to METH degradation and alter the interaction mechanism with the target organic contaminants. This result was in accordance with a previous study that the degradation rate of methyl paraben in UV/PS system was the highest at near neutral pH and decreased at pH 9 and pH 11 [30]. The pseudo-first order reaction rate constants of METH decreased most rapidly in the pH range of 7.0–9.0. The base may catalyze the hydrolysis of persulfate molecule to form the hydroperoxide intermediate (HO2-), which subsequently reacts with another persulfate molecule by single electron transfer generating a SO4-% (Eq. (13)) [13].
3.3. Effect of initial PS dosage The amount of PS dosage affects the removal efficiency of the oxidation process. METH degradation was significantly increased by the increasing PS dosage from 0 to 800 μM, and the reaction rate constants were enhanced from 0.005 min−1 to 1.14 min−1 (Fig. S5a). Degradation rates of METH were proportional to PS dosage in the UV/ PS system, which was ascribed to the generation of more reactive species with more oxidant amount [25]. As shown in Fig. S5b, the kobs and decomposed PS exhibited a linear relationship with a coefficient R2 of 0.98, which was consistent with the degradation of carbamazepine and sulfamethazine at different initial PS dosage in the UV/PS system [26]. It was reported that the scavenging of the excessive PS and the recombination of SO4-% could take place with an excessive amount of PS through Eqs. (10) and (11) [27,28]. In the present study this negative effect was not observed, probably because the maximum PS dosage (800 μM) was not high enough to inhibit the degradation.
Fig. 2. (a) The bimolecular reaction rate constant of %OH with METH. [METH]0=[BA]0 = 100 μg/ L, [H2O2]0 = 2 mM, pH = 7, temperature = 25 ± 1 °C; (b) The bimolecular reaction rate constant of SO4−% with METH. [METH]0 = [BA]0 = 100 μg/L, [PS]0 = 400 μM, [TBA] = 50 mM, pH = 7, temperature = 25 ± 1 °C.
34
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
It may explain the slight increase of METH degradation rate as pH increased from 9.0–11.0. The above result suggested that SO4-% was dominant at acidic and neutral pH, and %OH was dominant under alkaline condition, which was consistent with the previous studies [31,32]. The speciation of METH may lead to its different decomposition performance in aqueous solution at different pH. With a pKa value of 10.3, METH was presented in its protonated form at pH 3, pH 5 and pH 7 (Fig. S6). Competition kinetics at different pH was conducted to assess whether the two species of METH showed different reactivities towards SO4−% and %OH (Fig. S7). The obtained rate constants are presented in Table S3. Overall, the bimolecular rate constants of %OH and SO4-% with protonated form of METH were higher than those with deprotonated form, which was in agreement with the above discussion.
SO4 • + OH S2 O82 + 2H2 O
SO24 +•OH k = 7.3 × 107M 1s OH
(12)
1
(13)
2SO4 + SO4 • + O2 • + 4H+
Fig. 5. Effect of Cl− on METH degradation in UV/PS [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
3.5. Effect of natural water constituents
3.5.2. Cl− Remarkable suppressive effect of Cl− was observed even at low levels of Cl- on the METH degradation (Fig. 5). With the addition of 0.5 mM Cl-, the kobs dramatically dropped from 0.193 min-1 to 0.047 min-1. It was assumed that Cl- had a dual impact on the activated persulfate performance [34]. At low Cl- concentrations, SO4-% would react with Cl- to form Cl% (Eq. (18)) which has a high redox potential of 2.4 V and may decompose organic compound in a similar way as SO4-%. Whereafter, the generated Cl% might stimulate the propagation reaction, producing more SO4-% to enhance the degradation of target compound [35]. Nevertheless, at high Cl- concentrations, the formed Cl% could be further quenched by Cl- (Eq. (19)), and Cl2-% with a low redox potential of 1.36 V would be finally generated from SO4-% and lessen the overall reaction efficiency [36]. The inhibition effect was observed with the Clconcentrations ranging from 0 to 10 mM in the present study, likely because Cl2-% could not overwhelm the effect of SO4-% at low Cl- concentrations (0–10 mM).
HCO3−
3.5.1. Fig. 4 showed the METH degradation under different levels of HCO3− (0.5, 1, 2,5 10 mM), and the pH values were 7.2, 7.7, 8.1, 8.7, respectively. The kobs values decreased from 0.193 min-1 to 0.023 min-1 with increasing initial HCO3− concentration from 0 to 10 mM. HCO3− or CO32- could be transformed to HCO3%/CO3% (Eqs. (14)–(17)) [33] which is trivial and can be neglected for the degradation of most organic contaminants [14]. This result was inconsistent with the degradation of oxytetracycline in the presence of either HCO3− or CO32in the UV/H2O2 system, where a slight promotion effect was observed, likely because of the stronger oxidative ability of HCO3%/CO3% for oxytetracycline degradation [21].
SO4 • + HCO3 SO4 • + CO23
• OH + HCO3 • OH + CO32
CO•3 + H++SO24 CO3 •+SO24
k = 3.6 × 106M 1s
k = 6.5 × 106M 1s
CO3 •+HO
k = 4.2 × 108M 1s
1
(14) (15)
1
CO3 • + H2 O k = 8.5 × 106M 1s
1
1
system.
SO4 • + Cl
(16)
Cl• + Cl
(17)
Fig. 4. Effect of HCO3− on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
Cl•+SO24
k = 4.7 × 10 8M 1s
+Cl2 • k = 8.0 × 109M 1s
1
1
Fig. 6. Effect of NO3− on METH degradation in UV/PS [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
35
(18) (19)
system.
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
3.5.3. NO3− The effect of NO3− on METH degradation was shown in Fig. 6. With the addition of 0.5 mM and 10 mM of NO3− into the reaction system, the pseudo-first order rate constant kobs was reduced by nearly 2 and 10 times, respectively. Irradiated under UV light, NO3− could be converted to NO2% (1.03 V) and O-% which subsequently react with H2O to produce % OH through Eqs. (20) and (21) [37]. Moreover, SO4−% could convert NO3− to relatively low redox potential of NO3% (2.30 V) with a rate constant of 3.6 × 105 M-1 s-1 (Eq. (22) [38]). These produced active radicals would compete with SO4−% for METH degradation, which could consequently decrease the degradation rate of METH. An additional experiment was conducted in presence of NO3- alone (Fig. S8). Results showed the METH degradation was slightly enhanced (< 6%) with the addition of NO3- (10 mM), indicating negligible degradation of METH by %OH generated from NO3-. Furthermore, strong absorbance (200–240 nm) was observed with different concentration of NO3(0–10 mM, Fig. S9), indicating that other mechanism rather than UV filtering was responsible for the inhibition of UV/PS reaction by NO3-. This result was in accordance with the decomposition of 1H-benzotriazole in UV/PS system [39], during which strong absorbance (200–240 nm) was also found in the NO3- solution.
NO3 + h
O
•
+ H2 O
SO4 • + NO3
•
(20)
OH + OH
(21)
NO•2 + O •
NO•3+SO24
k = 3.6 × 105M 1s
1
competing with the target pollutant for oxidant species [43]. However, limited information was available on the interaction between NOM and SO4−%, which requird further investigation on the mechanism of strong inhibition effect from HA. 3.6. Degradation intermediates and pathway Degradation intermediates and products generated during the UV/ PS oxidation of METH were determined with UPLC/MS/MS under full scans and product ion scans. The mass spectra obtained before, during, and after METH degradation were compared to identify the intermediates, of which the structure was not determined due to the lack of standards. In this study, the structure of the transformation products was identified with the specific molecular ions and fragmentation patterns rather than direct comparison with corresponding standards. The mass spectra and proposed structures of the reaction products were depicted in Fig. S11 and Table S4, respectively. Two possible transformation pathways of METH under UV/PS were presented in Fig. 8 based on the proposed structures of six identified products. The proposed reaction mechanisms involved in the METH degradation include hydroxylation, dehydration, demethylation and deamination. Intermediate P1 (m/z = 166) was produced through hydroxylation (pathway 1) with the increment of 16 mass unit to METH (m/z = 150), which was consistent with the fragmentation patterns as previously described [10]. Intermediate P2 (m/z = 119) was formed via cleavage of −CH3 and -NH2CH3. The molecular mass of P3 (m/z = 101) was lower than that of P2 by 18 mass units, suggesting the dehydration of P2 into P3. Hydroxylation of the aromatic ring could take place through electrophilic attack by %OH and e− transfer from aromatic chemicals to the %OH to form radicals, providing %OH with adducts on hydrolysis [44,45]. High selective SO4−% could easily react with the aromatic rings via electron transfer. Nonselective %OH reacts with the aliphatic chain or aromatic rings through hydrogen abstraction or hydroxyl addition [44]. Radical quenching experiment indicated that both SO4−% and %OH were existing in the UV/PS system for METH removal, thus the two radicals could contribute to the hydroxylation process. Demethylation (pathway 2) resulted from the cleavage of methyl moiety led to the production of intermediate P4 (m/z = 135), which was consistent with the research reported by Fukahori et al [45]. SO4−% could effectively remove the alkyl groups in atrazine. Consequently, SO4−% was believed to trigger the demethylation via hydrogen abstraction in the methyl group. Further reactions of deamination may produce the intermediate P5 (m/z = 119), which was supposed to generate the product P6 (m/z = 91) through the cleavage of C2H4. The proposed intermediates could be further oxidized into smaller fragments via the subsequent attack of SO4−% and %OH radicals.
(22)
3.5.4. Humic acid As an inner filter, HA in the aqueous solution would greatly inhibit the UV transmittance for UV254 nm photons resulting in the reduction of SO4−% generation in UV/PS system [40]. HA may also decrease the reaction by competing for oxidant species with the target contaminants [41]. The METH degradation was obviously hindered with only a comparatively low concentration of HA (Fig. 7). Once the HA concentration was greater than 1 mM, the increase of inhibitory effect was neglected as the kobs slightly decreased from 0.021 min-1 to 0.016 min-1 with HA concentrations increasing from 2 to 20 mM. As an inner filter, HA could absorb UV irradiation (Fig. S10) and greatly decrease the UV transmittance for UV photons, resulting in the reduction of SO4−% generation in UV/PS system [42]. HA might also inhibit the reaction by
4. Conclusions The kinetics and degradation mechanism of METH by UV/PS oxidation were investigated in this study. The combination of UV and PS effectively eliminated METH with higher efficiency than direct UV photolysis or PS oxidation alone. The rate constant of METH degradation was proportional to the PS dosage and was high at neutral pH condition. The METH degradation was more attributed to SO4−% than % OH, favorable at neutral pH and inhibited by inorganic anions. Meanwhile, humic acid significantly suppressed METH degradation by the effect of UV absorption and free radical quenching. Possible transformation mechanisms involved in the METH degradation included hydroxylation, dehydration, demethylation and deamination. Results demonstrated that UV/PS was an efficient technique for the removal of METH, which provided a potential practical application for the destruction of other organic pollutants in natural water.
Fig. 7. Effect of HA on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
36
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
Fig. 8. Tentative transformation pathways of METH in the UV/PS system.
Acknowledgements
[11] T. Mackuľak, L. Birošová, I. Bodík, R. Grabic, A. Takáčová, M. Smolinská, A. Hanusová, J. Híveš, M. Gál, Zerovalent iron and iron(VI): effective means for the removal of psychoactive pharmaceuticals and illicit drugs from wastewaters, Sci. Total Environ. 539 (2016) 420–426. [12] S. Wacławek, H.V. Lutze, K. Grübel, V.V.T. Padil, M. Černík, D.D. Dionysiou, Chemistry of persulfates in water and wastewater treatment: a review, Chem. Eng. J. 330 (2017) 44–62. [13] L.W. Matzek, K.E. Carter, Activated persulfate for organic chemical degradation: a review, Chemosphere 151 (2016) 178–188. [14] Z. Wang, Y. Shao, N. Gao, N. An, Degradation kinetic of dibutyl phthalate (DBP) by sulfate radical- and hydroxyl radical-based advanced oxidation process in UV/ persulfate system, Sep. Purif. Technol. 195 (2017) 92–100. [15] P. Xie, J. Ma, W. Liu, J. Zou, S. Yue, X. Li, M.R. Wiesner, J. Fang, Removal of 2-MIB and geosmin using UV/persulfate: contributions of hydroxyl and sulfate radicals, Water Res. 69 (2015) 223–233. [16] R. Li, J. Kong, H. Liu, P. Chen, Y. Su, G. Liu, W. Lv, R. Li, J. Kong, H. Liu, Removal of indomethacin using UV-vis/peroxydisulfate: kinetics, toxicity, and transformation pathways, Chem. Eng. J. 331 (2017) 809–817. [17] J. Ye, P. Zhou, Y. Chen, H. Ou, J. Liu, C. Li, Q. Li, Degradation of 1H-benzotriazole using ultraviolet activating persulfate: mechanisms, products and toxicological analysis, Chem. Eng. J. 334 (2018) 1493–1501. [18] Z. Wang, Y. Chen, P. Xie, R. Shang, J. Ma, Removal of microcystis aeruginosa by UV-activated persulfate: performance and characteristics, Chem. Eng. J. 300 (2016) 245–253. [19] Y. Qian, X. Guo, Y. Zhang, Y. Peng, P. Sun, C. Huang, J. Niu, X. Zhou, J.C. Crittenden, Perfluorooctanoic acid degradation using UV–Persulfate process: modeling of the degradation and chlorate formation, Environ. Sci. Technol. 50 (2015) 772–781. [20] X. He, M. Pelaez, J.A. Westrick, K.E.O. Shea, A. Hiskia, T. Triantis, T. Kaloudis, M.I. Stefan, A.A. de la Cruz, D.D. Dionysiou, Efficient removal of microcystin-LR by UV-C/H2O2 in synthetic and natural water samples, Water Res. 46 (2012) 1501–1510. [21] Y. Liu, X. He, X. Duan, Y. Fu, D.D. Dionysiou, Photochemical degradation of oxytetracycline: influence of pH and role of carbonate radical, Chem. Eng. J. 276 (2015) 113–121. [22] C. Liang, C. Huang, N. Mohanty, R.M. Kurakalva, A rapid spectrophotometric determination of persulfate anion in ISCO, Chemosphere 73 (2008) 1540–1543. [23] L. Ismail, C. Ferronato, L. Fine, F. Jaber, J.M. Chovelon, Elimination of sulfaclozine from water with SO4-. radicals: evaluation of different persulfate activation methods, Appl. Catal. B Environ. 201 (2016) 573–581. [24] K. Yin, L. Deng, J. Luo, J. Crittenden, C. Liu, Y. Wei, L. Wang, Destruction of phenicol antibiotics using the UV/H2O2 process: kinetics, byproducts, toxicity evaluation and trichloromethane formation potential, Chem. Eng. J. 351 (2018) 867–877. [25] J.Y. Fang, C. Shang, Bromate formation from bromide oxidation by the UV/persulfate process, Environ. Sci. Technol. 46 (2012) 8976–8983.
This work was funded by the National Natural Science Foundation of China (41673120) and Beijing Natural Science Foundation (8173058). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jphotochem.2019.05. 009. References [1] C. Wei, K. Yi, G. Sun, J. Wang, Synthesis of novel sonocatalyst Er3+:YAlO3/Nb2O5 and its application for sonocatalytic degradation of methamphetamine hydrochloride, Ultrason. Sonochem. 42 (2018) 57–67. [2] L. Bijlsma, R. Serrano, C. Ferrer, I. Tormos, F. Hernández, Occurrence and behavior of illicit drugs and metabolites in sewage water from the Spanish Mediterranean coast (Valencia region), Sci. Total Environ. 487 (2014) 703–709. [3] A.J. Skees, K.S. Foppe, B. Loganathan, B. Subedi, Contamination profiles, mass loadings, and sewage epidemiology of neuropsychiatric and illicit drugs in wastewater and river waters from a community in the midwestern United States, Sci. Total Environ. 631 (2018) 1457–1464. [4] K.S. Foppe, D.R. Hammond-Weinberger, B. Subedi, Estimation of the consumption of illicit drugs during special events in two communities in Western Kentucky, USA using sewage epidemiology, Sci. Total Environ. 633 (2018) 249–256. [5] K. Li, P. Du, Z. Xu, T. Gao, X. Li, Occurrence of illicit drugs in surface waters in China, Environ. Pollut. 213 (2016) 395–402. [6] Y. Zhang, T. Zhang, C. Guo, J. Lv, Z. Hua, S. Hou, Y. Zhang, W. Meng, J. Xu, Drugs of abuse and their metabolites in the urban rivers of Beijing, China: occurrence, distribution, and potential environmental risk, Sci. Total Environ. 579 (2016) 305–313. [7] L.N. Kates, C.W. Knapp, H.E. Keenan, Acute and chronic environmental effects of clandestine methamphetamine waste, Sci. Total Environ. 493 (2014) 781–788. [8] R. Pal, M. Megharaj, K.P. Kirkbride, R. Naidu, Illicit drugs and the environment — a review, Sci. Total Environ. 463 (2013) 1079–1092. [9] A. Rodayan, P.A. Segura, V. Yargeau, Ozonation of wastewater: removal and transformation products of drugs of abuse, Sci. Total Environ. 487 (2014) 763–770. [10] C. Kuo, C. Lin, P.A.K. Hong, Photocatalytic degradation of methamphetamine by UV/TiO2 - kinetics, intermediates, and products, Water Res. 74 (2015) 1–9.
37
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al. [26] J. Deng, Y. Shao, N. Gao, S. Xia, C. Tan, S. Zhou, X. Hu, Degradation of the antiepileptic drug carbamazepine upon different UV-based advanced oxidation processes in water, Chem. Eng. J. 222 (2013) 150–158. [27] X. Lu, Y. Shao, N. Gao, J. Chen, Y. Zhang, H. Xiang, Y. Guo, Degradation of diclofenac by UV-activated persulfate process: kinetic studies, degradation pathways and toxicity assessments, Ecotoxicol. Environ. Safety 141 (2017) 139–147. [28] Y. Liu, X. He, Y. Fu, D.D. Dionysiou, Kinetics and mechanism investigation on the destruction of oxytetracycline by UV-254 nm activation of persulfate, J. Hazard. Mater. 305 (2016) 229–239. [29] C. Liang, Z.S. Wang, C.J. Bruell, Influence of pH on persulfate oxidation of TCE at ambient temperatures, Chemosphere 66 (2007) 106–113. [30] S. Dhaka, R. Kumar, M.A. Khan, K.J. Paeng, M.B. Kurade, S.J. Kim, B.H. Jeon, Aqueous phase degradation of methyl paraben using UV-activated persulfate method, Chem. Eng. J. 321 (2017) 11–19. [31] Y. Ji, C. Dong, D. Kong, J. Lu, Q. Zhou, Heat-activated persulfate oxidation of atrazine: implications for remediation of groundwater contaminated by herbicides, Chem. Eng. J. 263 (2015) 45–54. [32] J. Wang, S. Wang, Activation of persulfate (PS) and peroxymonosulfate (PMS) and application for the degradation of emerging contaminants, Chem. Eng. J. 334 (2017) 1502–1517. [33] Y. Yang, J.J. Pignatello, J. Ma, W.A. Mitch, Comparison of halide impacts on the efficiency of contaminant degradation by sulfate and hydroxyl radical-based advanced oxidation processes (AOPs), Environ. Sci. Technol. 48 (2014) 2344–2351. [34] Y. Deng, C.M. Ezyske, Sulfate radical-advanced oxidation process (SR-AOP) for simultaneous removal of refractory organic contaminants and ammonia in landfill leachate, Water Res. 45 (2011) 6189–6194. [35] C. Tan, N. Gao, S. Zhou, Y. Xiao, Z. Zhuang, Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes, Chem. Eng. J. 253 (2014) 229–236.
[36] O.S. Keen, N.G. Love, K.G. Linden, The role of effluent nitrate in trace organic chemical oxidation during UV disinfection, Water Res. 46 (2012) 5224–5234. [37] Y. Yang, J.J. Pignatello, J. Ma, W.A. Mitch, Comparison of halide impacts on the efficiency of contaminant degradation by sulfate and hydroxyl radical-based advanced oxidation processes (AOPs), Environ. Sci. Technol. 48 (2014) 2344–2351. [38] J. Ye, P. Zhou, Y. Chen, H. Ou, J. Liu, C. Li, Q. Li, Degradation of 1H-benzotriazole using ultraviolet activating persulfate: mechanisms, products and toxicological analysis, Chem. Eng. J. 334 (2017) 1493–1501. [39] B.T. Oh, Y.S. Seo, D. Sudhakar, J.H. Choe, S.M. Lee, Y.J. Park, M. Cho, Oxidative degradation of endotoxin by advanced oxidation process (O3/H2O2 & UV/H2O2, J. Hazard. Mater. 279 (2014) 105–110. [40] H.V. Lutze, S. Bircher, I. Rapp, N. Kerlin, R. Bakkour, M. Geisler, S.C. Von, T.C. Schmidt, Degradation of chlorotriazine pesticides by sulfate radicals and the influence of organic matter, Environ. Sci. Technol. 49 (2015) 1673–1680. [41] B. Darsinou, Z. Frontistis, M. Antonopoulou, I. Konstantinou, D. Mantzavinos, Sonoactivated persulfate oxidation of bisphenol a: kinetics, pathways and the controversial role of temperature, Chem. Eng. J. 280 (2015) 623–633. [42] R. Hazime, Q.H. Nguyen, C. Ferronato, A. Salvador, F. Jaber, J.M. Chovelon, Comparative study of imazalil degradation in three systems: UV/TiO2, UV/K2S2O8 and UV/TiO2/K2S2O8, Appl. Catal. B Environ. 144 (2014) 286–291. [43] F. Minisci, A. Citterio, C. Giordano, Electron-transfer processes: peroxydisulfate, a useful and versatile reagent in organic chemistry, Cheminform 14 (1983) 27–32. [44] S. Fukahori, T. Fujiwara, R. Ito, N. Funamizu, Photocatalytic decomposition of crotamiton over aqueous TiO2 suspensions: determination of intermediates and the reaction pathway, Chemosphere 89 (2012) 213–220. [45] J.A. Khan, X. He, N.S. Shah, H.M. Khan, E. Hapeshi, D. Fatta-Kassinos, D.D. Dionysiou, Kinetic and mechanism investigation on the photochemical degradation of atrazine with activated H2O2, S2O82− and HSO5-, Chem. Eng. J. 252 (2014) 393–403.
38
Contents lists available at ScienceDirect
Journal of Photochemistry & Photobiology A: Chemistry journal homepage: www.elsevier.com/locate/jphotochem
Removal of methamphetamine by UV-activated persulfate: Kinetics and mechanisms
T
Deming Gua,b, Changsheng Guoa, Jiapei Lva, Song Houa, Yan Zhanga, Qiyan Fengb, Yuan Zhanga, ⁎ Jian Xua, a b
State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing 100012, China School of Environment Science and Spatial Informatics, China University of Mining and Technology, Xuzhou 221116, China
ARTICLE INFO
ABSTRACT
Keywords: Methamphetamine Persulfate Sulfate radical Hydroxyl radical Advanced oxidation processes (AOP)
In this study, removal of methamphetamine (METH) was investigated by UV activated persulfate (PS), and the influence of key factors was evaluated. Results suggested that METH degradation followed pseudo-first order reaction kinetics. The combination of UV and persulfate (UV/PS) could completely degrade 100 μg/L of METH in 30 min with a PS dosage of 200 μM at pH 7. Both hydroxyl radical (%OH) and sulfate radical (SO4−%) were confirmed to contribute to the degradation of METH. The bimolecular reaction rate constants of METH with %OH and SO4−% were 7.91 × 109 and 3.29 × 109 M−1 s−1, respectively. The degradation rate constant of METH was proportional to the PS dosage (0–800 μM) and was high at neutral pH condition. The presence of inorganic anions significantly reduced METH degradation to different degrees, with the inhibitory effect order of Cl− > NO3− > HCO3−. The degradation efficiency of METH was suppressed by the presence of humic acid due to the effect of UV absorption and free radical quenching. The degradation intermediates and products were identified by UPLC-MS/MS and possible transformation pathways were proposed. Results suggested that the combination of UV/PS is a promising treatment technique for the removal of METH in the water environment.
1. Introduction Methamphetamine (METH) is a commonly consumed drug of abuse. According to World Drug Report 2018, METH contributes a large proportion to the synthetic drugs in the world illicit drugs market [1]. After being consumed, METH cannot be completely metabolized in the body, and the parent METH and its metabolites could be excreted into the environment via urine and feces. In the conventional wastewater treatment plants, the complete elimination of METH can rarely be achieved [2], leading to its frequent detection in different water bodies around the world [3,4]. For instance, a recent survey covering 49 lakes and 4 major rivers across China revealed that METH and ketamine were found with high levels and detection frequency [5]. In the urban rivers in Beijing, our previous study showed that METH was one of the most predominant drugs with detection frequency ranging from 65% to 100% [6]. Although METH was observed at trace concentrations in the natural aquatic ecosystem, it may pose negative effect to both humans and other organisms via long-term exposure [7,8]. Many measures have been attempted to treat METH from aqueous solution. Rodayan et al. [9] reported that ozonation was capable of
⁎
eliminating drugs of abuse including METH from wastewater, with the generation of some persistent intermediates. METH can be efficiently removed by TiO2 photocatalysis under UV365 irradiation, with a complete elimination of METH in 3 h with the dosage of TiO2 at 0.1 g/L [10]. Fenton and Fenton-like reactions were also effective to degrade psychoactive pharmaceuticals and illicit drugs to below the detection limit in the wastewater [11]. Activated persulfate oxidation is also one of the advanced oxidation processes (AOPs) that has drawn great interest because of the distinctive advantage of sulfate radical (SO4−%) with high stability, selectivity, redox potential (2.5–3.1 V) and applicability in wide pH ranges [12]. SO4−% could be produced from persulfate under the activated conditions of heat, light, transition metal, alkaline environment, activated carbon, electrochemical, complexing agent and other new types [13]. Of all the treatment techniques, UV/PS process is deemed as a costly, sustainable and environment-friendly method, which can generate SO4−% with a quantum yield of 0.7 mol E/s (Eq. (1)) [14], and other subsequent primary chemical reactions (Eqs. (2)–(5) [15].
S2 O82
Corresponding author. E-mail address: [email protected] (J. Xu).
https://doi.org/10.1016/j.jphotochem.2019.05.009 Received 1 March 2019; Received in revised form 20 April 2019; Accepted 5 May 2019 Available online 06 May 2019 1010-6030/ © 2019 Elsevier B.V. All rights reserved.
UV
2SO4 •
= 0.7 molE/s
(1)
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
SO4 • + OH
SO24 +•OH k = 6.5 × 107M 1s
SO4 • + H2 O
HSO24 +•OH k = 8.3M 1s
• OH+S2 O82
S2 O•8 +OH
• OH+SO•4
1
(3)
1
k = 1.4 × 107M 1s
HSO5 k = 1 × 1010M 1s
(2)
1
1
(4) (5)
The combination of UV and PS has been confirmed to destruct a range of contaminants, such as indomethacin [16], 1H-benzotriazole [17], microcystis aeruginosa [18] and perfluorooctanoic acid [19]. To our best knowledge, no information is available on the elimination of METH by UV/PS. The aim of the present study is to (1) investigate the degradation and mechanism of METH in the UV/PS system under various experimental condition, (2) determine the bimolecular reaction rate constants of METH with %OH and SO4−% through competitive kinetics, and (3) identify the oxidation intermediates. 2. Materials and methods Fig. 1. Removal of METH under different oxidation treatment. [METH]0 = 100 μg/L, initial pH = 7.0, PS dosage was 200 μM, temperature = 25 ± 1 °C.
2.1. Materials METH was obtained from Cerilliant Corporation (Round Rock, TX, USA), with detailed information listed in supplementary materials Table S1. HPLC grade acetonitrile and methanol (MeOH) were obtained from Fisher Scientific (Poole, UK). Sodium persulfate (PS, 98.0%), formic acid (FA, ≥98%), tert-butyl-alcohol (TBA) and benzoic acid (BA) were obtained from Sigma-Aldrich (Bellefonte, USA). Analytical grade hydrogen peroxide (H2O2, 30%, v/v), sodium bicarbonate (NaHCO3, 99.7%), sodium chloride (NaCl, ≥99.0%), sodium nitrate (NaNO3, ≥99.5%), humic acid (HA), Sulfuric acid (H2SO4, ≥98%) and sodium hydroxide (NaOH, ≥99.5%) were obtained from Sinopharm Chemical Reagent Co., Ltd. (Beijing, China) and used without further purification. All solutions were prepared using Milli-Q water produced by an ultrapure water system (Millipore, MA, USA).
USA). The mobile phase comprised of 0.1% FA in Milli-Q water (phase A) and acetonitrile (phase B), and the flow rate was kept at 450 μL min−1. Initial proportion of phase B (10%) was held for 0.5 min, and linearly increased to 45% within 1.8 min, then increased to 95% at 1.9 min, held for 1.0 min, reverted to the initial condition (10%) within 0.1 min, and kept for 1.5 min. 5 μL of analytes were injected, with the column temperature at 40 °C. The positive ion multiple reaction monitoring (MRM) mode was adopted to record chromatograms. The capillary voltage was set at 0.5 kV. Desolvation temperature was 400 °C and source temperature was 150 °C. Nitrogen was used as the desolvation and nebulizing gas. Detailed UPLC-MS/MS parameters were given in Table S2. A UV–vis spectrophotometer was used to measure the concentration of residual PS [22].
2.2. Experimental section
3. Results and discussion
The experiments were conducted in a photo-reactor (Fig. S1) with an effective volume of 50 mL (XPA-7, Xujiang Machinery Factory, Nanjing, China). A low-pressure mercury lamp (11 W, Philips Co., China) was placed in the quartz sleeve. A filter has been used to allow only 254 nm wavelength transmission. To keep the UV irradiation stable, the lamp was warmed up before experiment. The UV fluence rate of 0.1 mW cm−2 was determined by three different methods [20]. The quantum yield of METH at 254 nm was determined to be 0.013 Einstein-1 [21]. Appropriate volumes of freshly prepared METH and PS stock solutions were added to achieve 50 mL reaction solution. The homogeneous mixtures were then stirred thoroughly with electromagnetic stirrers (300 rpm). The reaction solution was moved under UV irradiation to initiate the reaction at pH 7.0 and room temperature. Samples were taken at specific time intervals and quenched immediately using the same volume of methanol and passed through 0.22 μm nylon filter prior to further analysis. The initial pH values were adjusted with H2SO4 or NaOH solution (0.1 M). MeOH or TBA was added respectively as scavengers for %OH or SO4-% to investigate the reactive species involved in the UV/PS reaction system.
3.1. Removal of METH by UV, PS and UV/PS The removal of METH by UV, PS and UV/PS oxidation was illustrated in Fig. 1. METH degradation well fit the pseudo-first order model (R2 > 0.95, Eq. (6)). Negligible removal of METH was observed by direct UV irradiation or PS oxidation alone, with the same degradation rate constant of 0.005 min−1. In contrast, METH was completely removed within 30 min under UV/PS treatment. The generation of %OH and SO4-% in the UV/PS system was likely responsible for the enhanced elimination of METH. The pseudo-first order rate constant of METH degradation by UV/PS was calculated to be 0.193 min−1. Slight changes (less than 15%) were observed for the concentration of remaining PS during PS oxidation alone or UV/PS system. The mineralization of METH was characterized by removal of TOC, which achieved 44% within 30 min under UV/PS treatment (Fig. S2).
In
[METH] [METH]0
=
kobs t
(6)
where [METH] is the concentration of METH at time t (min), [METH]0 is the initial concentration, and kobs is the pseudo-first order rate constant (min−1).
2.3. Analytical methods The concentration of METH was quantified by an ultra-performance liquid chromatography tandem mass spectrometry (UPLC-MS/MS) consisting of a Waters ACQUITY liquid chromatography and an Xevo T_QS triple quadrupole mass spectrometer (ESI-MS/MS, 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,
3.2. Determination of bimolecular reaction rate constants Radical scavengers were used to identify the predominant radical species. The results indicated METH degradation was attributed to both 33
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
OH and SO4−% oxidation (detailed description in Text S1). The bimolecular reaction rate constants of METH with %OH and SO4−% were determined by competition kinetics at neutral pH. Benzoic acid (BA) was used as the probe chemical, and the constant reaction rates of %OH and SO4−% with BA were 5.9 × 109 M-1 s-1 and 1.2 × 109 M-1 s-1, respectively [23]. Less than 10% of METH and BA were degraded under UV irradiation alone, thus direct photolysis was negligible. The METH degradation kinetics in UV/PS system could be calculated with the following Eq. (7):
%
d[METH] = kSO4 • dt
METH [METH][SO4
•
] + k•OH
METH [METH][•OH]
(7)
TBA (50 mM) was added into the solution for scavenging hydroxyl radical in the UV/PS system. METH and BA were degraded only by SO4−%. The bimolecular reaction rate constant of METH with SO4−% could be obtained by Eq. (8):
kSO4 •
METH
= kSO4 •
BA
In([METH]t /[METH]0 ) In([BA]t /[BA]0 )
(8)
Fig. 3. Effect of initial pH on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
As above, the bimolecular reaction rate constant of METH with OH at pH 7 in UV-254 nm/H2O2 system can be obtained with the following Eq. (9): %
k•OH
METH
= k•OH
In([METH]t /[METH]0 ) BA In([BA]t /[BA]0 )
SO4 • + S2 O28
(9)
SO4 • + SO4 •
The plot of ln([METH]0/[METH]t)/ln([BA]0/[BA]t) yielded a straight line with a slope of 1.34 and 2.74, as presented in Fig. 2. The bimolecular reaction rate constants of METH with %OH and SO4−% at pH 7 were 7.91 × 109 and 3.29 × 109 M-1 s-1, respectively. It is reported that the reaction rate constants of organic compounds by %OH varied from 106 to 1010 M-1 s-1 [24], suggesting the rapid reaction with %OH.
S2 O8 • + SO24
S2 O82
k = 1.2 × 106M 1s
k = 4 × 10 8M 1s
1
1
(10) (11)
3.4. Effect of solution pH Solution pH may affect the degradation rates because of the difference of electronic distribution in molecules. As shown in Fig. 3, METH degradation in UV/PS system was dependent on pH, with the kobs values following the order of pH 7.0 (0.193 min−1) > pH 3.0 (0.125 min−1) > pH 5.0 (0.089 min−1) > pH 11.0 (0.032 min−1) > pH 9.0 (0.021 min−1). Overall, the reaction rates under acidic and natural conditions were higher than that under base conditions, which was likely attributed to the low activation energy of SO4-% under neutral pH condition [29]. Moreover, SO4-% was able to react with OH- to produce lower redox potential of %OH under alkaline conditions (Eq. (12) [28], which might lessen the contribution of SO4-% to METH degradation and alter the interaction mechanism with the target organic contaminants. This result was in accordance with a previous study that the degradation rate of methyl paraben in UV/PS system was the highest at near neutral pH and decreased at pH 9 and pH 11 [30]. The pseudo-first order reaction rate constants of METH decreased most rapidly in the pH range of 7.0–9.0. The base may catalyze the hydrolysis of persulfate molecule to form the hydroperoxide intermediate (HO2-), which subsequently reacts with another persulfate molecule by single electron transfer generating a SO4-% (Eq. (13)) [13].
3.3. Effect of initial PS dosage The amount of PS dosage affects the removal efficiency of the oxidation process. METH degradation was significantly increased by the increasing PS dosage from 0 to 800 μM, and the reaction rate constants were enhanced from 0.005 min−1 to 1.14 min−1 (Fig. S5a). Degradation rates of METH were proportional to PS dosage in the UV/ PS system, which was ascribed to the generation of more reactive species with more oxidant amount [25]. As shown in Fig. S5b, the kobs and decomposed PS exhibited a linear relationship with a coefficient R2 of 0.98, which was consistent with the degradation of carbamazepine and sulfamethazine at different initial PS dosage in the UV/PS system [26]. It was reported that the scavenging of the excessive PS and the recombination of SO4-% could take place with an excessive amount of PS through Eqs. (10) and (11) [27,28]. In the present study this negative effect was not observed, probably because the maximum PS dosage (800 μM) was not high enough to inhibit the degradation.
Fig. 2. (a) The bimolecular reaction rate constant of %OH with METH. [METH]0=[BA]0 = 100 μg/ L, [H2O2]0 = 2 mM, pH = 7, temperature = 25 ± 1 °C; (b) The bimolecular reaction rate constant of SO4−% with METH. [METH]0 = [BA]0 = 100 μg/L, [PS]0 = 400 μM, [TBA] = 50 mM, pH = 7, temperature = 25 ± 1 °C.
34
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
It may explain the slight increase of METH degradation rate as pH increased from 9.0–11.0. The above result suggested that SO4-% was dominant at acidic and neutral pH, and %OH was dominant under alkaline condition, which was consistent with the previous studies [31,32]. The speciation of METH may lead to its different decomposition performance in aqueous solution at different pH. With a pKa value of 10.3, METH was presented in its protonated form at pH 3, pH 5 and pH 7 (Fig. S6). Competition kinetics at different pH was conducted to assess whether the two species of METH showed different reactivities towards SO4−% and %OH (Fig. S7). The obtained rate constants are presented in Table S3. Overall, the bimolecular rate constants of %OH and SO4-% with protonated form of METH were higher than those with deprotonated form, which was in agreement with the above discussion.
SO4 • + OH S2 O82 + 2H2 O
SO24 +•OH k = 7.3 × 107M 1s OH
(12)
1
(13)
2SO4 + SO4 • + O2 • + 4H+
Fig. 5. Effect of Cl− on METH degradation in UV/PS [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
3.5. Effect of natural water constituents
3.5.2. Cl− Remarkable suppressive effect of Cl− was observed even at low levels of Cl- on the METH degradation (Fig. 5). With the addition of 0.5 mM Cl-, the kobs dramatically dropped from 0.193 min-1 to 0.047 min-1. It was assumed that Cl- had a dual impact on the activated persulfate performance [34]. At low Cl- concentrations, SO4-% would react with Cl- to form Cl% (Eq. (18)) which has a high redox potential of 2.4 V and may decompose organic compound in a similar way as SO4-%. Whereafter, the generated Cl% might stimulate the propagation reaction, producing more SO4-% to enhance the degradation of target compound [35]. Nevertheless, at high Cl- concentrations, the formed Cl% could be further quenched by Cl- (Eq. (19)), and Cl2-% with a low redox potential of 1.36 V would be finally generated from SO4-% and lessen the overall reaction efficiency [36]. The inhibition effect was observed with the Clconcentrations ranging from 0 to 10 mM in the present study, likely because Cl2-% could not overwhelm the effect of SO4-% at low Cl- concentrations (0–10 mM).
HCO3−
3.5.1. Fig. 4 showed the METH degradation under different levels of HCO3− (0.5, 1, 2,5 10 mM), and the pH values were 7.2, 7.7, 8.1, 8.7, respectively. The kobs values decreased from 0.193 min-1 to 0.023 min-1 with increasing initial HCO3− concentration from 0 to 10 mM. HCO3− or CO32- could be transformed to HCO3%/CO3% (Eqs. (14)–(17)) [33] which is trivial and can be neglected for the degradation of most organic contaminants [14]. This result was inconsistent with the degradation of oxytetracycline in the presence of either HCO3− or CO32in the UV/H2O2 system, where a slight promotion effect was observed, likely because of the stronger oxidative ability of HCO3%/CO3% for oxytetracycline degradation [21].
SO4 • + HCO3 SO4 • + CO23
• OH + HCO3 • OH + CO32
CO•3 + H++SO24 CO3 •+SO24
k = 3.6 × 106M 1s
k = 6.5 × 106M 1s
CO3 •+HO
k = 4.2 × 108M 1s
1
(14) (15)
1
CO3 • + H2 O k = 8.5 × 106M 1s
1
1
system.
SO4 • + Cl
(16)
Cl• + Cl
(17)
Fig. 4. Effect of HCO3− on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
Cl•+SO24
k = 4.7 × 10 8M 1s
+Cl2 • k = 8.0 × 109M 1s
1
1
Fig. 6. Effect of NO3− on METH degradation in UV/PS [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
35
(18) (19)
system.
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
3.5.3. NO3− The effect of NO3− on METH degradation was shown in Fig. 6. With the addition of 0.5 mM and 10 mM of NO3− into the reaction system, the pseudo-first order rate constant kobs was reduced by nearly 2 and 10 times, respectively. Irradiated under UV light, NO3− could be converted to NO2% (1.03 V) and O-% which subsequently react with H2O to produce % OH through Eqs. (20) and (21) [37]. Moreover, SO4−% could convert NO3− to relatively low redox potential of NO3% (2.30 V) with a rate constant of 3.6 × 105 M-1 s-1 (Eq. (22) [38]). These produced active radicals would compete with SO4−% for METH degradation, which could consequently decrease the degradation rate of METH. An additional experiment was conducted in presence of NO3- alone (Fig. S8). Results showed the METH degradation was slightly enhanced (< 6%) with the addition of NO3- (10 mM), indicating negligible degradation of METH by %OH generated from NO3-. Furthermore, strong absorbance (200–240 nm) was observed with different concentration of NO3(0–10 mM, Fig. S9), indicating that other mechanism rather than UV filtering was responsible for the inhibition of UV/PS reaction by NO3-. This result was in accordance with the decomposition of 1H-benzotriazole in UV/PS system [39], during which strong absorbance (200–240 nm) was also found in the NO3- solution.
NO3 + h
O
•
+ H2 O
SO4 • + NO3
•
(20)
OH + OH
(21)
NO•2 + O •
NO•3+SO24
k = 3.6 × 105M 1s
1
competing with the target pollutant for oxidant species [43]. However, limited information was available on the interaction between NOM and SO4−%, which requird further investigation on the mechanism of strong inhibition effect from HA. 3.6. Degradation intermediates and pathway Degradation intermediates and products generated during the UV/ PS oxidation of METH were determined with UPLC/MS/MS under full scans and product ion scans. The mass spectra obtained before, during, and after METH degradation were compared to identify the intermediates, of which the structure was not determined due to the lack of standards. In this study, the structure of the transformation products was identified with the specific molecular ions and fragmentation patterns rather than direct comparison with corresponding standards. The mass spectra and proposed structures of the reaction products were depicted in Fig. S11 and Table S4, respectively. Two possible transformation pathways of METH under UV/PS were presented in Fig. 8 based on the proposed structures of six identified products. The proposed reaction mechanisms involved in the METH degradation include hydroxylation, dehydration, demethylation and deamination. Intermediate P1 (m/z = 166) was produced through hydroxylation (pathway 1) with the increment of 16 mass unit to METH (m/z = 150), which was consistent with the fragmentation patterns as previously described [10]. Intermediate P2 (m/z = 119) was formed via cleavage of −CH3 and -NH2CH3. The molecular mass of P3 (m/z = 101) was lower than that of P2 by 18 mass units, suggesting the dehydration of P2 into P3. Hydroxylation of the aromatic ring could take place through electrophilic attack by %OH and e− transfer from aromatic chemicals to the %OH to form radicals, providing %OH with adducts on hydrolysis [44,45]. High selective SO4−% could easily react with the aromatic rings via electron transfer. Nonselective %OH reacts with the aliphatic chain or aromatic rings through hydrogen abstraction or hydroxyl addition [44]. Radical quenching experiment indicated that both SO4−% and %OH were existing in the UV/PS system for METH removal, thus the two radicals could contribute to the hydroxylation process. Demethylation (pathway 2) resulted from the cleavage of methyl moiety led to the production of intermediate P4 (m/z = 135), which was consistent with the research reported by Fukahori et al [45]. SO4−% could effectively remove the alkyl groups in atrazine. Consequently, SO4−% was believed to trigger the demethylation via hydrogen abstraction in the methyl group. Further reactions of deamination may produce the intermediate P5 (m/z = 119), which was supposed to generate the product P6 (m/z = 91) through the cleavage of C2H4. The proposed intermediates could be further oxidized into smaller fragments via the subsequent attack of SO4−% and %OH radicals.
(22)
3.5.4. Humic acid As an inner filter, HA in the aqueous solution would greatly inhibit the UV transmittance for UV254 nm photons resulting in the reduction of SO4−% generation in UV/PS system [40]. HA may also decrease the reaction by competing for oxidant species with the target contaminants [41]. The METH degradation was obviously hindered with only a comparatively low concentration of HA (Fig. 7). Once the HA concentration was greater than 1 mM, the increase of inhibitory effect was neglected as the kobs slightly decreased from 0.021 min-1 to 0.016 min-1 with HA concentrations increasing from 2 to 20 mM. As an inner filter, HA could absorb UV irradiation (Fig. S10) and greatly decrease the UV transmittance for UV photons, resulting in the reduction of SO4−% generation in UV/PS system [42]. HA might also inhibit the reaction by
4. Conclusions The kinetics and degradation mechanism of METH by UV/PS oxidation were investigated in this study. The combination of UV and PS effectively eliminated METH with higher efficiency than direct UV photolysis or PS oxidation alone. The rate constant of METH degradation was proportional to the PS dosage and was high at neutral pH condition. The METH degradation was more attributed to SO4−% than % OH, favorable at neutral pH and inhibited by inorganic anions. Meanwhile, humic acid significantly suppressed METH degradation by the effect of UV absorption and free radical quenching. Possible transformation mechanisms involved in the METH degradation included hydroxylation, dehydration, demethylation and deamination. Results demonstrated that UV/PS was an efficient technique for the removal of METH, which provided a potential practical application for the destruction of other organic pollutants in natural water.
Fig. 7. Effect of HA on METH degradation in UV/PS system. [METH]0 = 100 μg/L, [PS]0 = 200 μM, temperature = 25 ± 1 °C.
36
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al.
Fig. 8. Tentative transformation pathways of METH in the UV/PS system.
Acknowledgements
[11] T. Mackuľak, L. Birošová, I. Bodík, R. Grabic, A. Takáčová, M. Smolinská, A. Hanusová, J. Híveš, M. Gál, Zerovalent iron and iron(VI): effective means for the removal of psychoactive pharmaceuticals and illicit drugs from wastewaters, Sci. Total Environ. 539 (2016) 420–426. [12] S. Wacławek, H.V. Lutze, K. Grübel, V.V.T. Padil, M. Černík, D.D. Dionysiou, Chemistry of persulfates in water and wastewater treatment: a review, Chem. Eng. J. 330 (2017) 44–62. [13] L.W. Matzek, K.E. Carter, Activated persulfate for organic chemical degradation: a review, Chemosphere 151 (2016) 178–188. [14] Z. Wang, Y. Shao, N. Gao, N. An, Degradation kinetic of dibutyl phthalate (DBP) by sulfate radical- and hydroxyl radical-based advanced oxidation process in UV/ persulfate system, Sep. Purif. Technol. 195 (2017) 92–100. [15] P. Xie, J. Ma, W. Liu, J. Zou, S. Yue, X. Li, M.R. Wiesner, J. Fang, Removal of 2-MIB and geosmin using UV/persulfate: contributions of hydroxyl and sulfate radicals, Water Res. 69 (2015) 223–233. [16] R. Li, J. Kong, H. Liu, P. Chen, Y. Su, G. Liu, W. Lv, R. Li, J. Kong, H. Liu, Removal of indomethacin using UV-vis/peroxydisulfate: kinetics, toxicity, and transformation pathways, Chem. Eng. J. 331 (2017) 809–817. [17] J. Ye, P. Zhou, Y. Chen, H. Ou, J. Liu, C. Li, Q. Li, Degradation of 1H-benzotriazole using ultraviolet activating persulfate: mechanisms, products and toxicological analysis, Chem. Eng. J. 334 (2018) 1493–1501. [18] Z. Wang, Y. Chen, P. Xie, R. Shang, J. Ma, Removal of microcystis aeruginosa by UV-activated persulfate: performance and characteristics, Chem. Eng. J. 300 (2016) 245–253. [19] Y. Qian, X. Guo, Y. Zhang, Y. Peng, P. Sun, C. Huang, J. Niu, X. Zhou, J.C. Crittenden, Perfluorooctanoic acid degradation using UV–Persulfate process: modeling of the degradation and chlorate formation, Environ. Sci. Technol. 50 (2015) 772–781. [20] X. He, M. Pelaez, J.A. Westrick, K.E.O. Shea, A. Hiskia, T. Triantis, T. Kaloudis, M.I. Stefan, A.A. de la Cruz, D.D. Dionysiou, Efficient removal of microcystin-LR by UV-C/H2O2 in synthetic and natural water samples, Water Res. 46 (2012) 1501–1510. [21] Y. Liu, X. He, X. Duan, Y. Fu, D.D. Dionysiou, Photochemical degradation of oxytetracycline: influence of pH and role of carbonate radical, Chem. Eng. J. 276 (2015) 113–121. [22] C. Liang, C. Huang, N. Mohanty, R.M. Kurakalva, A rapid spectrophotometric determination of persulfate anion in ISCO, Chemosphere 73 (2008) 1540–1543. [23] L. Ismail, C. Ferronato, L. Fine, F. Jaber, J.M. Chovelon, Elimination of sulfaclozine from water with SO4-. radicals: evaluation of different persulfate activation methods, Appl. Catal. B Environ. 201 (2016) 573–581. [24] K. Yin, L. Deng, J. Luo, J. Crittenden, C. Liu, Y. Wei, L. Wang, Destruction of phenicol antibiotics using the UV/H2O2 process: kinetics, byproducts, toxicity evaluation and trichloromethane formation potential, Chem. Eng. J. 351 (2018) 867–877. [25] J.Y. Fang, C. Shang, Bromate formation from bromide oxidation by the UV/persulfate process, Environ. Sci. Technol. 46 (2012) 8976–8983.
This work was funded by the National Natural Science Foundation of China (41673120) and Beijing Natural Science Foundation (8173058). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jphotochem.2019.05. 009. References [1] C. Wei, K. Yi, G. Sun, J. Wang, Synthesis of novel sonocatalyst Er3+:YAlO3/Nb2O5 and its application for sonocatalytic degradation of methamphetamine hydrochloride, Ultrason. Sonochem. 42 (2018) 57–67. [2] L. Bijlsma, R. Serrano, C. Ferrer, I. Tormos, F. Hernández, Occurrence and behavior of illicit drugs and metabolites in sewage water from the Spanish Mediterranean coast (Valencia region), Sci. Total Environ. 487 (2014) 703–709. [3] A.J. Skees, K.S. Foppe, B. Loganathan, B. Subedi, Contamination profiles, mass loadings, and sewage epidemiology of neuropsychiatric and illicit drugs in wastewater and river waters from a community in the midwestern United States, Sci. Total Environ. 631 (2018) 1457–1464. [4] K.S. Foppe, D.R. Hammond-Weinberger, B. Subedi, Estimation of the consumption of illicit drugs during special events in two communities in Western Kentucky, USA using sewage epidemiology, Sci. Total Environ. 633 (2018) 249–256. [5] K. Li, P. Du, Z. Xu, T. Gao, X. Li, Occurrence of illicit drugs in surface waters in China, Environ. Pollut. 213 (2016) 395–402. [6] Y. Zhang, T. Zhang, C. Guo, J. Lv, Z. Hua, S. Hou, Y. Zhang, W. Meng, J. Xu, Drugs of abuse and their metabolites in the urban rivers of Beijing, China: occurrence, distribution, and potential environmental risk, Sci. Total Environ. 579 (2016) 305–313. [7] L.N. Kates, C.W. Knapp, H.E. Keenan, Acute and chronic environmental effects of clandestine methamphetamine waste, Sci. Total Environ. 493 (2014) 781–788. [8] R. Pal, M. Megharaj, K.P. Kirkbride, R. Naidu, Illicit drugs and the environment — a review, Sci. Total Environ. 463 (2013) 1079–1092. [9] A. Rodayan, P.A. Segura, V. Yargeau, Ozonation of wastewater: removal and transformation products of drugs of abuse, Sci. Total Environ. 487 (2014) 763–770. [10] C. Kuo, C. Lin, P.A.K. Hong, Photocatalytic degradation of methamphetamine by UV/TiO2 - kinetics, intermediates, and products, Water Res. 74 (2015) 1–9.
37
Journal of Photochemistry & Photobiology A: Chemistry 379 (2019) 32–38
D. Gu, et al. [26] J. Deng, Y. Shao, N. Gao, S. Xia, C. Tan, S. Zhou, X. Hu, Degradation of the antiepileptic drug carbamazepine upon different UV-based advanced oxidation processes in water, Chem. Eng. J. 222 (2013) 150–158. [27] X. Lu, Y. Shao, N. Gao, J. Chen, Y. Zhang, H. Xiang, Y. Guo, Degradation of diclofenac by UV-activated persulfate process: kinetic studies, degradation pathways and toxicity assessments, Ecotoxicol. Environ. Safety 141 (2017) 139–147. [28] Y. Liu, X. He, Y. Fu, D.D. Dionysiou, Kinetics and mechanism investigation on the destruction of oxytetracycline by UV-254 nm activation of persulfate, J. Hazard. Mater. 305 (2016) 229–239. [29] C. Liang, Z.S. Wang, C.J. Bruell, Influence of pH on persulfate oxidation of TCE at ambient temperatures, Chemosphere 66 (2007) 106–113. [30] S. Dhaka, R. Kumar, M.A. Khan, K.J. Paeng, M.B. Kurade, S.J. Kim, B.H. Jeon, Aqueous phase degradation of methyl paraben using UV-activated persulfate method, Chem. Eng. J. 321 (2017) 11–19. [31] Y. Ji, C. Dong, D. Kong, J. Lu, Q. Zhou, Heat-activated persulfate oxidation of atrazine: implications for remediation of groundwater contaminated by herbicides, Chem. Eng. J. 263 (2015) 45–54. [32] J. Wang, S. Wang, Activation of persulfate (PS) and peroxymonosulfate (PMS) and application for the degradation of emerging contaminants, Chem. Eng. J. 334 (2017) 1502–1517. [33] Y. Yang, J.J. Pignatello, J. Ma, W.A. Mitch, Comparison of halide impacts on the efficiency of contaminant degradation by sulfate and hydroxyl radical-based advanced oxidation processes (AOPs), Environ. Sci. Technol. 48 (2014) 2344–2351. [34] Y. Deng, C.M. Ezyske, Sulfate radical-advanced oxidation process (SR-AOP) for simultaneous removal of refractory organic contaminants and ammonia in landfill leachate, Water Res. 45 (2011) 6189–6194. [35] C. Tan, N. Gao, S. Zhou, Y. Xiao, Z. Zhuang, Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes, Chem. Eng. J. 253 (2014) 229–236.
[36] O.S. Keen, N.G. Love, K.G. Linden, The role of effluent nitrate in trace organic chemical oxidation during UV disinfection, Water Res. 46 (2012) 5224–5234. [37] Y. Yang, J.J. Pignatello, J. Ma, W.A. Mitch, Comparison of halide impacts on the efficiency of contaminant degradation by sulfate and hydroxyl radical-based advanced oxidation processes (AOPs), Environ. Sci. Technol. 48 (2014) 2344–2351. [38] J. Ye, P. Zhou, Y. Chen, H. Ou, J. Liu, C. Li, Q. Li, Degradation of 1H-benzotriazole using ultraviolet activating persulfate: mechanisms, products and toxicological analysis, Chem. Eng. J. 334 (2017) 1493–1501. [39] B.T. Oh, Y.S. Seo, D. Sudhakar, J.H. Choe, S.M. Lee, Y.J. Park, M. Cho, Oxidative degradation of endotoxin by advanced oxidation process (O3/H2O2 & UV/H2O2, J. Hazard. Mater. 279 (2014) 105–110. [40] H.V. Lutze, S. Bircher, I. Rapp, N. Kerlin, R. Bakkour, M. Geisler, S.C. Von, T.C. Schmidt, Degradation of chlorotriazine pesticides by sulfate radicals and the influence of organic matter, Environ. Sci. Technol. 49 (2015) 1673–1680. [41] B. Darsinou, Z. Frontistis, M. Antonopoulou, I. Konstantinou, D. Mantzavinos, Sonoactivated persulfate oxidation of bisphenol a: kinetics, pathways and the controversial role of temperature, Chem. Eng. J. 280 (2015) 623–633. [42] R. Hazime, Q.H. Nguyen, C. Ferronato, A. Salvador, F. Jaber, J.M. Chovelon, Comparative study of imazalil degradation in three systems: UV/TiO2, UV/K2S2O8 and UV/TiO2/K2S2O8, Appl. Catal. B Environ. 144 (2014) 286–291. [43] F. Minisci, A. Citterio, C. Giordano, Electron-transfer processes: peroxydisulfate, a useful and versatile reagent in organic chemistry, Cheminform 14 (1983) 27–32. [44] S. Fukahori, T. Fujiwara, R. Ito, N. Funamizu, Photocatalytic decomposition of crotamiton over aqueous TiO2 suspensions: determination of intermediates and the reaction pathway, Chemosphere 89 (2012) 213–220. [45] J.A. Khan, X. He, N.S. Shah, H.M. Khan, E. Hapeshi, D. Fatta-Kassinos, D.D. Dionysiou, Kinetic and mechanism investigation on the photochemical degradation of atrazine with activated H2O2, S2O82− and HSO5-, Chem. Eng. J. 252 (2014) 393–403.
38