Degradation of atrazine in heterogeneous Co3O4 activated peroxymonosulfate oxidation process: Kinetics, mechanisms, and reaction pathways

Accepted Manuscript Degradation of atrazine in heterogeneous Co3O4 activated peroxymonosulfate oxidation process: Kinetics, mechanisms, and reaction pathways Yan Fan, Yuefei Ji, Guanyu Zheng, Junhe Lu, Deyang Kong, Xiaoming Yin, Quansuo Zhou PII: DOI: Reference:

S1385-8947(17)31360-8 http://dx.doi.org/10.1016/j.cej.2017.08.020 CEJ 17471

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

12 April 2017 25 July 2017 6 August 2017

Please cite this article as: Y. Fan, Y. Ji, G. Zheng, J. Lu, D. Kong, X. Yin, Q. Zhou, Degradation of atrazine in heterogeneous Co3O4 activated peroxymonosulfate oxidation process: Kinetics, mechanisms, and reaction pathways, Chemical Engineering Journal (2017), doi: http://dx.doi.org/10.1016/j.cej.2017.08.020

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Degradation of atrazine in heterogeneous Co 3O4 activated peroxymonosulfate oxidation process: Kinetics, mechanisms, and reaction pathways Yan Fana, Yuefei Jia, Guanyu Zhenga, Junhe Lua*, Deyang Kongb*, Xiaoming Yina, Quansuo Zhoua

a

Department of Environmental Science and Engineering, Nanjing Agricultural University, Nanjing, 210095, China

b

Nanjing Institute of Environmental Science, Ministry of Environmental Protection of PRC, Nanjing, 210042, China

*Corresponding author: E-mail: [email protected]; [email protected]

Telephone: (86) 25-84395164; Fax: (86) 25-84395210

1

Abstract: Herbicide atrazine (ATZ) has caused great environmental concern due to its long-term application in agriculture and persistence nature. This study examined the degradation of ATZ in heterogeneously activated peroxymonosulfate (PMS) oxidation processes, using Co3O4 as the catalyst, finding that high PMS concentrations and near neutral pH (pH 6.0) were beneficial for ATZ degradation. ATZ degradation rate was influenced by functional groups present on the Co3 O4 surface at varying pH. Complete removal of 20 µM ATZ was achieved in 15 min, with 2.0 mM PMS and 0.4 g/L Co3O4 at pH 6.0. Minimal cobalt leaching occurred during reaction, shown by a maximum dissolved Co concentration (0.06 mg/L) found at pH 3.0 and decreasing with increasing pH. During the reaction, Co3 O4 showed high potential for reusability. Structural properties of the pristine and used Co3 O4 catalysts were characterized by scanning electron microscope; X-ray diffraction; and X-ray photoelectron spectroscopy, with no changes observed post-reaction. A total of 7 intermediate products of ATZ were detected by liquid chromatography-tandem mass spectrometry, with quantification of intermediate products allowing the contribution of each pathway for ATZ degradation, to be accessed. Transformation pathways including dealkylation, dechlorination-hydroxylation and alkylic-oxidation were proposed for catalytic decomposition of ATZ in the Co3 O4/PMS system.

Keywords: atrazine; peroxymonosulfate; Co3O4 ; sulfate radical; mechanisms

2

1 Introduction

Hydroxyl radical (•OH)-based advanced oxidation processes (AOPs) have been established as effective methods for the removal and mineralization of various organic contaminants in water. More recently, sulfate radical (SO4•−) has received notable attention as an alternative for •OH, due to a comparable, or even higher oxidation potential (2.5-3.1 V) than •OH (1.8-2.7 V)[1]; a longer half-life time in aqueous solution (30-40 µs)[2]; and a wider operational pH range [3, 4]. Previous studies have indicated that SO4•− reacts more selectively and efficiently with certain organic compounds via electron transfer mechanisms[1, 5], for example, perfluorinated carboxylic acids can be degraded by SO4•− but are inert to •OH[6]. SO4•− can be generated by activation of peroxymonosulfate (PMS) or peroxydisulfate (PDS) by UV[7]; base[8]; heat[9]; and transition metals[10-12]. Among these methods, transition metal activation is generally easier to operate and more cost-effective than the energy-based approaches[12]. Comparably, PMS is more easily activated by transition metals due to its asymmetrical molecular structure[12]. A number of transition metals such as Co, Fe, and Cu can be used to activate PMS for SO4•− production[10-12, 14], among them, the Co2+ was proven to be most effective[11].

Co (II) is highly soluble at environmentally relevant conditions. Co (II) presents primarily as Co2+ at pH below 9.4 and Co(OH)2 at pH above 9.4. Its speciation against the change of pH is shown in Figure S1. The mechanism of Co2+ activation of PMS is depicted in Fig.S2, showing that the key step is the formation of CoOH+[15], with SO4•− produced during the conversion of CoOH+ to CoO+, prior to conversion to Co3+. Meanwhile, Co3+ is reduced back to Co2+ by PMS and SO5•− generated as the by-product, with this step maintaining the catalytic potential of cobalt[16]. It has been reported that Co2+/PMS system is effective in the degradation 3

of a wide range of pollutants, including persistent organic pollutants (POPs)[11]; azo dyes[16]; estrogens[17]; and herbicides[11, 18], since the generation of SO4•− by Co2+ catalyzed PMS decomposition was initially proposed by Ball et al.[14] However, a major disadvantage of this system is that homogeneous Co2+/PMS system results in the discharge of dissolved Co into the environment. Co, a toxic and potentially carcinogenic heavy metal, is regulated globally[19, 20], with Co discharge increasing operational costs as well as causing serious environmental concerns. To overcome this disadvantage of homogeneous Co2+/PMS system, heterogeneous Co catalysts such as CoO, CoO2, Co2O3, and Co3 O4, have been examined for PMS activation. Activation of PMS by CoO and Co3O4 was assessed by Anipsitakis et al. for 2,4-dichlorophenol (2,4-DCP) degradation[21], showing that these particles could activate PMS efficiently with low levels of dissolved cobalt leaching; Shi et al. found that Co3O4 activated PMS systems could efficiently degrade Orange II in water[22]; while Wang et al. successfully applied this system for phenol removal[23]. Overall, the low level of cobalt leaching and highly efficient PMS activation, make Co3O4 a promising catalyst with low environmental risk.

Atrazine (2-chloro-4-ethylamino-6-isopropylamino-s-triazine, ATZ) is a herbicide used for the control of broadleaf weeds and grasses[24]. ATZ has been classed as an endocrine disrupting chemical and a possible human carcinogen[25, 26]. Due to these concerning associations, some European countries have banned the use of atrazine, although it can still be detected globally in soils, surface waters and ground waters, due to its high solubility, persistence and long-term history of use [27]. The U. S. EPA established an ATZ limit in drinking water of no more than 3 µg/L[28, 29]. Therefore, technologies that can effectively remove ATZ have received significant attention. To date, various methods have been investigated for ATZ removal from soils and waters, 4

for example, Khandarkhaeva et al. demonstrated that solar-enhanced Fenton-like processes showed efficient ATZ removal in water[30]. Luo et al. found ATZ can be successfully degraded by UV/H2O2, UV/HSO5 - and UV/S2O82- systems[31]. Ji et al. demonstrated that homogenous Co2+/PMS system efficiently degraded ATZ[18]. However, the removal of ATZ using a heterogeneous Co3O4/PMS system has not yet been reported.

The purpose of this study was to investigate the degradation of ATZ in a heterogeneous Co3O4/PMS system. The degradation kinetics, Co leaching, and the reusability of Co3O4 were systematically examined, to assess the feasibility of practical application of this method. Both pristine Co3O4 and used Co3O4 were characterized by scanning electron microscope (SEM), X-ray diffraction (XRD) and X-ray photoelectron spectroscopy (XPS). In addition, a series of intermediates and products were identified via liquid chromatography-tandem mass spectrometry (HPLC-MS/MS), allowing detailed mechanisms and transformation pathways for ATZ degradation in Co3O4/PMS systems to be proposed based on structural analysis of degradation products.

2 Materials and methods

2.1 Chemicals

Atrazine (99.9%), potassiummonopersulfate triple salt (KHSO5·0.5KHSO4·0.5K2SO4; ≥ 47% KHSO5 basis), methyl orange (MO), cobalt sulfate heptahydrate (CoSO4·7H2O), deethyl-atrazine (DEA, 99%), deisopropyl-atrazine (DIA, 95.4%), deethyl-hydroxy-atrazine (DEHA, 98.7%), deethyl-deisopropyl-atrazine (DEIA, 98.3%), deethyl-deisopropyl-hydroxy-atrazine (DEIHA, 99.6%), HPLC grade methanol and t-butanol (TBA) were all purchased from Aladdin Chemistry

5

Co., Ltd. (Shanghai, China). Cobalt oxide (Co3O4 ，99.9%) was obtained from Xiya Research Center (Shangdong, China). All other reagents were of analytical grade or higher and were used as received. The stock solutions were prepared by dissolving the reagents in Milli-Q water (>18 MΩ/cm) prepared using a Millipore Milli-Q system. Stock solutions of ATZ (100 µM); PMS (100 mM); MO (776 µM); CoSO4·7H2O (20 mM) were papered and stored in the dark at 4 oC, with all stock solutions used within 1 month.

Surface water was sampled from Tai Lake (Wuxi, China). Groundwater was sampled from a local well (Nanjing, China). Water quality parameters for both samples were given in the supporting information (Table S1). The samples were filtered through 0.45 µM membrane to remove colloids before an appropriate amount of ATZ was added to achieve a concentration of 20 µM for each of them.

2.2 Kinetic study

Batch reactions were performed in 30 mL glass vials at room temperature. Reactions were initiated by the addition of 0.4 g/L Co3O4 to solutions containing 20 µM ATZ and 2 mM PMS. Solution pH was adjusted to the desired value using 0.1 mM H2SO4 or 0.1 mM NaOH unless otherwise stated. The change of pH during the reaction course was within 0.2 unit. At predetermined time intervals, 0.5 mL sample aliquots were withdrawn, filtered through a 0.22 µm membrane and immediately mixed with excessive methanol (0.5 mL) to quench any reactions involving •OH or SO4•− [18]. Pre-experiment demonstrated that such treatment could effectively suppress the degradation of ATZ. The reaction was performed at pH 3.0, 4.0, 5.0, 6.0, 7.0, 8.0, 9.0, and 10.0, to evaluate the effects of pH on system performance. Control experiments were performed concurrently, under identical conditions except in the absence of either PMS or Co3O4. 6

To distinguish the contribution of •OH and SO4•−, removal of ATZ was performed in the presence of 200 mM ethanol or TBA at pH 6. All samples were stored at 4 oC prior to analysis. Residual ATZ of the samples were determined using a Hitachi L-2000 HPLC equipped with a photo diode array detector (DAD) and C18 reverse phase column (Hitachi LaChrom, 5µM × 250 mm × 4.6 mm). The isocratic elution used for the mobile phase consisted of 70% methanol and 30% water, with a flow rate of 1.0 mL/min, a detection wavelength of 222 nm and an injection volume of 10 µL.

2.3 Analysis of PMS concentration

Decomposition of PMS during the reactions described above was also explored. Residual PMS in the solution was quantified according to the method developed by Zou et al.[32] This method utilizes the decolorization of methyl orange (MO) by SO4 •− generated during PMS activation by Co2+, where a linear relationship can be established between the degree of decolorization of MO and PMS concentration. Briefly, 1 mL MO stock solution (776 µM) and 8 mL Milli-Q water were combined in a 30 mL glass vial, with 1 mL of CoSO4 stock solution (20 mM) and 0.1 mL of reaction sample containing PMS added sequentially. Following a 1 minute reaction period, the solution absorbance was measured at 464 nm (the characteristic absorption wavelength of MO) using a VARIAN Cary 50 UV-vis Spectrophotometer. Blank solutions were prepared using the sample in the absence of PMS or Co2+.

2.4 Co leaching test

The same experimental setup and conditions were used to quantify the Co leaching into solution, where at pre-determined time points (15minutes and 2 days) 8.0 mL supernatant was

7

withdrawn, filtered through a 0.22 µm membrane and transferred to a 10 mL glass vial. The concentration

of

dissolved

Co

in

samples

was

inductively coupled plasma-optical emission spectrometry

analyzed

using

(ICP-OES),

an

with

Agilent the

700

operating

parameters as follows: RF power 1.50 kW; sampling depth 8.0 mm; plasma gas flow 15 L/min; auxiliary gas flow 0.9 L/min; nebulizer flow 0.8 L/min; with the measured isotope

238.892

Co. The

detection limit of this approach was 0.2 µg/L.

2.5 Recycling of Co3O4

A 10 mL solution containing 20 µM ATZ and 2 mM PMS at pH 5.0, was combined with 4 mg Co3O4 in a centrifuge tube. Following a 10 minute reaction period, the mixture was centrifuged for 10 min at 10000 rpm and supernatant was collected and filtered through a 0.22 µm membrane. The residual ATZ concentration in the samples was quantified by HPLC, while the remaining Co3O4 was rinsed 5 times with Milli-Q water prior to the addition of 10 mL solution for the 2nd round of reaction. This process was repeated, resulting in a total of 9 repeated runs using the same Co3O4 material. Following the 9th run, the used Co3O4 particles were collected via centrifugation, washed with deionized water to remove impurities and oven dried at 50 oC to a constant weight.

2.6 Characterization of pristine and used Co3 O4 material

The morphology of pristine and used Co3O4 material were observed using a Quanta 400 FEG scanning electron microscope (SEM) operated at an acceleration voltage of 20 kV. A crystal graphic study of pristine and used Co3O4 was carried out using a D8-ADVANCE powder diffractometer at a scanning rate of 10 °/min in the 2θ range of 5-80 ° with Cu-Kα radiation (20

8

kV and 40 mA) at room temperature. X-ray photoelectron spectroscopy (XPS) measurement was performed using a Thermo ESCALAB 250 XI spectrometer with Al-Kα (E = 1361 eV) as the X-ray source, measuring the atomic composition of surfaces of pristine and used Co3O4 material. XPS spectra were corrected using the C 1s line at 284.6 eV.

2.7 Analysis of ATZ degradation intermediates and products

To characterize ATZ transformation products formed, reactions were performed at pH 6.0 with a relatively high initial ATZ concentration of 40 µM, with a PMS concentration of 2 mM and Co3O4 0.4 g/L. At set time intervals, 0.5 mL aliquots were removed and the reaction immediately quenched by mixing with 0.5 mL methanol. The quenched sample was then analyzed using a HPLC-MS/MS system consisting of an Agilent 1200 series HPLC coupled to an Agilent 6410 triple quadrupole mass spectrometer (Agilent Technologies, USA). MS analysis was carried out in positive mode using an electrospray ionization (ESI) source, with the instrument parameters as follows: capillary voltage 4.0 kV; fragmentor 135 V; desolvation by nitrogen gas (≥ 99.995%) at a flow rate of 10 L/min; temperature 350 oC; nebulizer pressure 40 psi; with nitrogen (≥ 99.999%) used as collision gas. The mass analyzer was operated in full scan mode (m/z range 100-500) to identify reaction products and once possible products were identified, selected ion scan (SIM) was performed. Selected intermediates were quantified by comparing the peak areas with authentic standards.

3 Results and discussion

3.1 Kinetics of ATZ degradation

No ATZ degradation could be observed in the presence of PMS or Co3O4 alone, while ATZ 9

was efficiently degraded in Co3O4/PMS systems, with a higher PMS concentration resulting in more efficient ATZ degradation as shown in Fig. 1. Activation of PMS by cobalt generates SO4•−[11]. However, •OH can also be formed by reaction between SO4•− and OH−, particularly under neutral and basic conditions (R1) [1]. As the source of free radicals, a higher PMS concentration is expected to result in higher steady-state concentrations of free radicals. Experimental data illustrated in Fig. 1 show that the degradation of ATZ was accelerated as PMS concentration increased. ATZ removal increased from 16% to 99% with increase of the initial PMS concentration from 0.1 to 2.0 mM, over a 15 minute period. In addition, the removal of ATZ appeared to be first order to the concentration of ATZ (Fig.1), and the rate can be described by Eq 1. 

 SO∙  + OH → SO  + ∙ OH

−

[]



k = 6.5 × 107 M-1s-1

= k  [ATZ]

(R 1)

(Eq 1)

k  = ∑ k  α [PMS]

(Eq 2)

where [ATZ] is the concentration of ATZ at certain reaction time; kobs is the pseudo first-order rate constant for ATZ degradation. Such a phenomenon indicates that PMS was in great excess and its

consumption during the investigated time period was negligible. The insert in Fig.1 shows a linear relationship can be established between kobs and PMS concentration (Eq 2). Considering it was the free radicals that caused the removal of ATZ, the dada reveal that the steady-state concentrations of free radicals were proportional to PMS concentration. In Eq 2, ki is the second order rate constant for ATZ reaction with radical species i; and αi is the yield of radical i (e.g., SO4•− and •OH) generated from PMS activation.

10

Fig.1 Degradation of ATZ at different concentrations of PMS. Insert is kobs versus PMS concentrations. [ATZ]0= 20 µM; Co3 O4 = 0.4 g/L; pH 6.0; 25 ◦ C. Error bars represent the standard deviation of 3 replicates. In order to distinguish the contribution of SO4•− and •OH to the reaction, removal of ATZ at pH 6.0 was examined in the presence of ethanol and TBA, respectively. TBA reacts with •OH at a second-order rate constant 1000-folder higher than SO4•− (R2 and 3), while both SO4•− and •OH react with ethanol at similar rate (R4 and 5) [3]. Thus, ethanol quenches the reactions mediated by both SO4•− and •OH, while TBA quenches •OH selectively. It was found that the pseudo-first order kinetic constant of ATZ removal was reduced from 0.173 min-1 to 0.029 min-1 by adding 200 mM TBA, while completely quenched by 200 mM ethanol. The data indicate that the removal of ATZ was attributed to the reaction with SO4•− and •OH; while •OH even played a more important role than SO4•− at pH 6.0.

TBA + ∙ OH → products

k = (3.8~7.6) × 108 M-1s-1

(R2)

TBA + SO∙  → products

k = (4.0~9.1) × 108 M-1s-1

(R3)

EtOH + ∙ OH → products

k = (1.2~2.8) × 108 M-1s-1

(R4)

11

EtOH + SO∙  → products

k = (1.6~7.7) × 108 M-1s-1

(R5)

3.2 Effects of pH

Solution pH has a significant effect on the performance of Co3O4/PMS systems in pollutants degradation, by affecting the surface charge of Co3 O4 particles as well as the speciation of pollutants and oxidants. ATZ degradation experiments were performed at a range of pH, with results presented in Fig.2. It was observed that the degradation rate of ATZ increased significantly when pH increased from 3.0 to 6.0. For example, in 15 minutes only 31% of ATZ was removed at pH 3.0, while almost complete removal was achieved at pH 6.0, with a rate constant of 0.2941 min-1 at pH 6.0, 14-fold higher than that at pH 3.0 (0.0208 min-1). The degradation of ATZ slowed down dramatically when pH was increased to above 6.0, with only 60% and 55% of ATZ degraded in 15 min at pH 7.0 and 8.0, respectively. Overall, the data demonstrate that the Co3 O4/PMS system showed optimal performance at near-neutral and environmentally-relevant pH, which is a significant advantage for practical applications.

Fig.2 The pseudo-first order rate constants for ATZ degradation and PMS decomposition (insert) at varying pH. [ATZ]0= 20 µM; [PMS]0= 2.0 mM; Co3O4 = 0.4 g/L; 25 ◦ C. Error bars represent the standard deviation of 3 replicates. 12

ATZ has a pKa of 1.68 [33]. At the pH range (3-6) investigated in this study, it presented as neutral form and its speciation on the reaction can be ignored. However, solution pH can affect both the functional groups on the surface of Co3O4 and the speciation of PMS, which in turn affects the interactions that occur between PMS and Co3O4, and consequentially free radical generation. This may partially explain alterations in ATZ degradation rates at varying pH. As shown in Fig.2, the degradation rate of ATZ was consistent with the decomposition rate of PMS (insert in Fig.2) at pH 3.0-6.0. Generally, when metal oxide particles are present in water, hydrolysis on the surface occurs[34]. When the solution pH is below pHpzc (isoelectric point, 6.3 for Co3O4), the particle becomes protonated, forming ≡CoOH+ (R.6), as opposed to ≡CoOOHbeing formed (R.8) which would result in the particle being negatively charged. Therefore, at pH < 6.3, Co3O4 maintained a net positive charge, facilitating its interaction with negatively charged PMS due to electrostatic attraction and resulting in favorable decomposition of PMS. According to Anipsitakis et al., CoOH+ plays a key role in homogeneous PMS activation[11], similarly, it is presumed that ≡CoOH+ was the key activator in Co3O4/PMS system. However, with reduction in pH, ≡CoOH+ is further protonated to form ≡CoOH22+ (R.7), inhibiting PMS decomposition. As seen in Fig.2 (insert), the first order rate constant for PMS decomposition decreased from 0.0173 to 0.0021 min-1 as pH reduced from 6.0 to 3.0. This effect was likely to be responsible for the decreased degradation rate of ATZ at pH 3.0 and 4.0. ≡ CoO + H- →≡ CoOH-

(pH < pHpzc)

≡ CoOH - + H - →≡ CoOH

≡ CoO + OH →≡ CoOOH 

(R. 6)

(R. 7)

(pH > pHpzc)

13

(R. 8)

0≡ Co - + HSO + SO∙ + OH . ↔≡ Co

(R. 9)

≡ Co0- + HSO + SO∙ . →≡ Co . + H

(R. 10)

When pH increased to above 6.3, ≡CoOOH- dominates and particles become negatively charged, resulting in electrostatic repulsive force preventing interaction between PMS and Co3O4. This resulted in the reduction in both PMS decomposition and ATZ degradation, as illustrated in Fig.3. However, it was also observed that PMS decomposition rate increased with an increase in pH from 7 to 10 (insert in Fig.2), which can be explained by the increased level of SO4•− consumption by OH-(R.1)[1, 35, 36], facilitating the decomposition of PMS. Reaction between SO4•− and OH- generates •OH and although ATZ shows a similar reactivity toward SO4•− (2.2~3.5×109 M-1s-1) and •OH (2.4~3.0×109 M-1s-1) at acidic conditions [37,38], •OH is less potent at neutral and basic conditions [12]. As •OH becomes the increasingly dominant from pH 7.0 to 10.0, the rate of ATZ degradation decreases. Furthermore, basic conditions are unfavorable for the reduction of Co3+ to Co2+(R.9 and 10), therefore inhibiting the generation of free radicals[39,40].

3.3 Co3O4 recycling and Co leaching

Recycling of Co3O4 is essential for practical application. Anipsitakis et al. found that Co3O4 maintained both excellent catalytic activity and stability following multiple cycled uses[21]. Nine repeated cycles of use were performed for the Co3O4 used in the present study. As shown in Fig.3a, the level of ATZ removal in 10 minutes decreased from 59 % to 30 % following 5 cycles of reuse, indicating that there was a reduction in the catalytic capability of Co3O4. However, ATZ degradation rate recovered gradually in consecutive runs. In the 9th cycle, 42% of ATZ was degraded within 10 minutes. While further studies are still required to optimize reusability, overall,

14

the data in Fig.3a demonstrates that Co3O4 has potential for efficient and effective reuse.

Fig.3 (a) Recycling of Co3O4 in Co3O4/PMS system; (b) Comparison of ATZ degradation in Co2+/PMS and Co3O4/PMS systems. [ATZ]0= 20 µM; [PMS]0= 2.0 mM; Co3O4 = 0.4 g/L; Co2+ = 0.06 mg/L; pH 6.0; 25 ◦ C. Error bars represents the standard deviation of 3 replicates.

Overall, Co leaching was observed at very low levels in the Co3 O4/PMS system investigated, with the concentration of dissolved Co presented in Table.1. An exception was that an appreciable Co leaching was observed at pH 3.0, with dissolved Co concentration reaching 0.06 mg/L in 15 minutes, although no further leaching was detected in the following 2 days of reaction. It is of note, that this level of Co discharge was significantly lower than that observed by Anipsitakis et al. in Co3O4 /PMS systems used for the degradation of chlorophenols[21]. At neutral

15

and alkaline conditions, dissolved Co concentrations were consistently lower than those at acidic conditions, with values below 0.02 mg/L following 15 minutes reaction time, below 0.04 mg/L following 2 days of reaction. In all cases, the concentrations of leached Co were far below the limitof 1.0 mg/L, established by the Environmental Quality Standards for Surface Water in China [20]. Previous study has shown that Co3O4 is less stable under acidic conditions[41], leading to high levels of Co leaching, while under alkaline conditions, Co2+ and Co3+ precipitate to form Co(OH)2 and Co(OH)3, respectively[11], resulting in a decrease of dissolved Co. Levels of Co leaching in controls with PMS absent are also listed in Table.1, showing no significant difference from the treatment systems.

Table.1 Concentrations of dissolved Co (mg/L) during ATZ degradation in Co3 O4/PMS system 15 min

2d

pH Control

Test

Control

Test

3

0.06

0.06

0.12

0.06

4

0.04

0.05

0.09

0.05

5

0.03

0.04

0.02

0.04

6

0.01

0.03

0.01

0.05

7

0.02

0.02

0

0.04

8

0.01

0.01

0

0.04

9

0.01

0.01

0

0.02

10

0

0

0

0

To evaluate the contribution of dissolved Co to the ATZ degradation, removal of ATZ in homogeneous Co2+/PMS system with 0.06 mg/L Co2+at pH 5.0 was examined. As shown in Fig.3b, the degradation rate of ATZ was significantly lower than that observed in Co3O4/PMS system. Therefore, it can be concluded that the degradation of ATZ in the Co3O4/PMS system was mainly 16

attributed to SO4•− generated by heterogeneous activation of PMS, rather than to the dissolved Co ions.

3.4 Characterization of Co3O4 material

To investigate the stability of the Co3 O4 catalyst, both pristine and used Co3O4 were characterized using SEM and XRD (Fig. 4), showing both the pristine and used Co3O4 particles have a similar spherical morphology, with diameters of approximately 5.0 µm. XRD patterns for both pristine and used Co3O4 particles (Fig. 5) displayed crystalline structures with eight typical broad characteristic peaks (2θ:19.0o, 31.2o, 36.8o, 38.5o, 44.8o, 55.7o, 59.3o, 65.2o). In addition, the diffraction peaks for both pristine and used Co3O4 match the standard diffraction data for Co3O4 (JCPDS No. 43-1003), showing that the used catalyst is still composed of pure Co3O4 particles and that the structure of Co3O4 was not changed after reaction. As compared to pristine Co3 O4, the XRD peak intensity for used Co3O4 was strengthened slightly due to crystal reunion during the reaction. Overall, Co3O4 kept a high level of stability during reactions, as demonstrated by the similarities in morphological and structural properties between pristine and used Co3O4 particles.

a

b

Fig.4 SEM images of the pristine (a) and used (b) Co3 O4 particles

17

Fig.5 XRD patterns of pristine and used Co3O4 particles

XPS was also utilized to investigate surface composition and elemental speciation of both pristine and used Co3 O4 particles, with the XPS spectra of Co3O4 surface metals and oxygen species shown in Fig.6. Each spectrum has two sharp peaks with binding energies at 779.8 and 795.0 eV, corresponding to Co 2p3/2 and Co 2p1/2, respectively[42].The peaks present at 781.0 and 796.1 eV can be attributed to Co2+ species, while the peaks at 779.7 and 794.7eV can be attributed to Co3+ species[43]. Therefore, Co2+/Co3+ ratios on the surfaces of pristine and used Co3O4 were established as 2.01 and 2.07, respectively. The very slight difference of Co2+/Co3+ ratio implies that Co2+ was mostly regenerated following Co3+ reduction, maintaining the outstanding activation potential of the Co3O4 catalyst. Two O 1s peaks can be observed with binding energies at 530.0 (Olattice, lattice oxygen) and 531.7 eV (Oadsorbed, surface adsorbed oxygen)[44]. Although Ren et al. suggest that accompanying the reduction of Co3+ to Co2+, Olattice is converted to dissolved O2 and Oadsorbed is transferred to ≡CoOH+[45], the Olattice/Oadsorbed ratio on the surface of used Co3 O4 particle (1.87) was only marginally higher than that of pristine Co3O4 particle (1.74) (Fig. 6b). Thus, the balance among Co2+/Co3+ and Olattice/Oadsorbed made the outstanding activation potential of Co3O4 catalyst.

18

Fig.6 XPS spectra of Co 2p (a) and O 1s (b) core levels for pristine and used Co3O4 particles

3.5 Removal of ATZ in real water samples

Degradation of ATZ in a surface water and a groundwater samples was examined to investigate the influence of water matrix on the performance of Co3O4/PMS system. Both water samples are slightly basic with pH 7.47 and 7.75 for the surface and ground water, respectively. But the surface water contains significantly higher total organic carbon (TOC, 13.69 mg/L) than the groundwater (4.05 mg/L). As it is demonstrated in Fig. 7, ATZ degradation rate decreased in the real water samples but the reactions were still pseudo-first order. The rate constants were 0.042 and 0.024 min-1 for the groundwater and surface water, respectively, in contrast to the value of 0.068 min-1 in pure water with pH 7.0 as the control (0.076 min-1). Such inhibition of degradation

19

was probably due to the scavenging effects of various organic and inorganic constituents in waters, especially the organic contents. Thus, high PMS dose is required in order to obtain satisfying removal of ATZ in real waters of high organic content, and the reaction conditions should be carefully optimized.

Fig. 7 Removal of ATZ in real water samples by Co3O4 /PMS. [ATZ]0 = 20 µM; Co3O4 = 0.4 g/L; 25◦ C; water quality parameters are given in Table S1. Error bars represent the standard deviation of 3 replicates.

3.6 Products identification and reaction pathways

Degradation products were analyzed by HPLC-ESI-MS/MS. Molecular structures of products were proposed according to MS/MS fragmentation patterns with comparison to available standards and literature data. A total of 7 intermediate products were identified, namely 2-chloro-4-acetamido-6-isopropylamino-1,3,5-trazine

(CAIT,

m/z

230);

2-hydroxy-4-acetamido-6-isopropylamino-1,3,5-trazine (HAIT, m/z 212); deethyl-atrazine (DEA, m/z 188); deisopropylatrazine (DIA, m/z 174); deethylhydroxyatrazine (DEHA, m/z 170); deethyldeisopropylatrazine (DEIA, m/z 146); deethyldeisopropylhydroxyatrazine (DEIHA, m/z 128)[18], with the time dependent formation of these intermediates presented in Fig.8. 20

Fig.8 Evolution of ATZ degradation products in Co3O4/PMS system: (a) dealkylation and dechlorination-hydroxylation products, (b) alkylchain oxidation products; (c) mass balance of quantified products. [ATZ]0= 40 µM; [PMS]0= 2.0 mM; Co3O4 = 0.4 g/L; pH 5.0; 25 ◦ C.

Dealkylation has frequently been observed during the oxidative degradation of ATZ by free radicals and other oxidants[47-50]. This is initialized by the formation of a carbon-center radical 21

generated by H-atom abstract due to the attack of SO4•- and/or HO• on the α-C adjacent to N atom. Subsequent oxidation of the carbon-center radical by O2 yielded a peroxide radical, which converted to a Shiff base by loss of a perhydroxyl radical (HO2•). Hydrolysis of the Shiff base leaded to the dealkylated products. In the present study, the cleavage of ethyl or isopropyl groups led to the formation of DEA or DIA, respectively. In accordance with Lutze et al., The dealkylated products appeared less reactive than the parent ATZ [49]. As shown in Fig.8a, DIA concentrations reached a temporal maximum of 1.97 µM in approximately 10 minutes, decreasing slowly subsequently. DEA formation was less rapid than DIA, reaching a maximum concentration of 2.12 µM by 40 minutes. Further dealkylation of either DIA or DEA resulted in the formation of DEIA.

Dechlorination-hydroxylation is another transformation pathway of ATZ transformation. It was an addition-elimination process by reaction with SO4•- or •OH, producing a radical cation at the carbon center. Sequent addition of H2O/OH- and elimination of a chlorine atom led to hydroxyl atrazine (HA). This intermediate has been observed in heat activated PDS oxidation of ATZ[9]. However, HA was not detected in this study, probably due to its fast dealkylation to form HAIT and DEHA.

CAIT was produced from alkyl oxidation of ethyl groups on ATZ molecule, which led to a carbinolamine intermeidate. Further oxidation of the carbinolamine by SO4•- or •OH generated CAIT. CAIT underwent dechlorination-hydroxylation subsequently resulting in HAIT formation. As shown in Fig.8b, CAIT increased steadily to a maximum concentration in 40 minutes, decreasing slowly thereafter, with a similar time-dependent evolution observed for HAIT. Unfortunately, CAIT and HAIT were only semi-quantified due to a lack of analytical standards and it is of note, that the hydrolysis of CAIT alkyl amino group could also lead to the formation of 22

DEA due to the loss of acetaldehyde[37].

DEHA could be formed through dechlorination-hydroxylation from DEA and hydrolysis of HAIT. As shown in Fig.8a, DEHA concentration increased to 0.12 µM in 30 minutes of reaction, gradually decreasing beyond this time, with the low level of DEHA production suggesting that it is not a dominant transformation pathway for ATZ.

Both DEIA and DEHA could be subsequently converted to DEIHA, the concentration of which increased continuously once detected with a total of 10.04 µM produced in 120 min of reaction. It is of note that no decrease was observed in DEIHA concentration over the course of the 120 minute reaction period, which may be attributed to its continuous formation and relative persistence to free radicals. A total of 85.6% reduction in ATZ was observed in 2 h, accompanied by the formation of DEA (4.3%); DIA (2.5%); DEIA (6.6%); DEHA (0.2%); and DEIHA (17.3%). While CAIT and HAIT were also formed at detectable levels, they were not quantified due to the lack of authentic standards. The sum of DEA, DIA, DEIA, DEHA and DEIHA formed increased to 15.63 µM in120minutes (Fig.8c), totally, c.a. 30.9 % of ATZ degradation was contributed to the formation of the 5 products. The remaining transformed ATZ might be converted to small molecule compounds, unidentified oxidation products, and unquantified CAIT and HAIT.

Based on the above analysis, a comprehensive reaction scheme involved in the oxidation of ATZ in Co3O4/PMS system can be proposed as depicted in Fig.9. The heterogeneous reactions were initiated by the hydroxylation of Co3O4 on surface metals resulting in the formation of ≡CoOH+ which was likely to be responsible for PMS activation. Then, HSO5− binds to Co3 O4 surfaces via hydrogen bonding (≡Co-O-H-HSO5−)[46] and ≡Co2+ species were oxidized to ≡Co3+ via electron transfer. Consequentially, SO4 •− was generated during the process of cleavage of O-H 23

and O-O bonds[46]. In Co3O4/PMS systems, SO4•− induced degradation of ATZ mainly proceeds through

three

pathways

including

dealkylation,

dechlorination-hydroxylation

and

alkylic-oxidation. The 3 pathways are overlapped and ultimately lead to DEIHA which was likely followed by the formation of cyanuric acid (not detected). According to Hooper et al., dealkylation products of ATZ such as DEA and DIA are less toxic than the parent ATZ[51]. Hydroxylated products generally show no toxicity effects to aquatic organisms. Thus, after the treatment of Co3O4/PMS, the hazardous effects associated with ATZ are expected to reduce significantly even complete mineralization is not achieved.

24

Fig.9 Proposed ATZ degradation pathways and mechanisms.

4 Conclusions

This study demonstrates that effective removal of ATZ from aqueous solutions can be achieved using Co3O4/PMS system, with the degradation rate of ATZ increasing with increased PMS concentrations. Furthermore, degradation efficiency is significantly related to solution pH, with the optimal removal observed at pH 6.0. Co3O4 exhibited good heterogeneous activity and 25

stability with low levels of cobalt leaching, suggesting the potential for good reusability. Results of SEM, XRD and XPS analysis indicate that no obvious changes were induced on the surface of Co3O4 following reaction. Possible degradation pathways of ATZ in heterogeneous Co3O4/PMS systems were proposed based on the molecular structures of identified degradation intermediates and

products.

The

main

pathways

of

ATZ

degradation

are

dealkylation,

dechlorination-hydroxylation and alkylic-oxidation. Several intermediate products were quantified to establish and evaluate the contribution of each pathway to ATZ degradation in Co3 O4/PMS systems. It is of note, the influence of various organic and inorganic constituents in water on the performance of the reaction system should be of concern and further research is required to optimize the Co3 O4/PMS system for practical application to ATZ degradation in wastewater treatment systems. In addition, immobilization of the fine Co3O4 particles to large carriers would be desirable in order to facilitate the separation of recycle of the catalyst.

Acknowledgments

This research was supported by the Fundamental Research Funds for Central Universities (KYZ201717), the National Science Foundation of China (51578294), and the Priority Academic Program Development (PAPD) of Jiangsu Higher Education Institute.

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33

Highlights •

Co3O4/PMS system is efficient for ATZ degradation.

•

The optimal ATZ removal was observed at pH 6.0.

•

Co3O4 presented a good reusability and limited Co leaching occurred during reaction.

•

ATZ was degraded via dealkylation, dechlorination, and alkylic oxidation pathways.

34