Nonradical degradation of microorganic pollutants by magnetic N-doped graphitic carbon: A complement to the unactivated peroxymonosulfate

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Nonradical degradation of microorganic pollutants by magnetic N-doped graphitic carbon: A complement to the unactivated peroxymonosulfate ⁎

Yong Fenga,b, Liyuan Zhangc, Zequn Yangc, Yiang Fanc, Kaimin Shihc, , Hailong Lid, Ying Liue, Deli Wue a

SCNU Environmental Research Institute, Guangdong Provincial Key Laboratory of Chemical Pollution and Environmental Safety & MOE Key Laboratory of Theoretical Chemistry of Environment, South China Normal University, Guangzhou 510006, China b School of Environment, South China Normal University, University Town, Guangzhou 510006, China c Department of Civil Engineering, The University of Hong Kong, Pokfulam, Hong Kong, China d School of Energy Science and Engineering, Central South University, Changsha 410083, China e State Key Laboratory of Pollution Control and Resources Reuse, School of Environmental Science & Engineering, Tongji University, Shanghai 200092, China

H I GH L IG H T S

G R A P H I C A L A B S T R A C T

N-doped graphitic carbon • Magnetic core–shell structures (MN-GCCSs) were synthesized.

had great catalytic re• MN-GCCSs activity for removal of microorganic contaminants.

were more reactive than • MN-GCCSs the commonly known activators of peroxymonosulfate.

Neither radicals nor O was the active • species. 94% of the bisphenol A was • Around degraded after 7th catalytic cycle. 1

2

A R T I C LE I N FO

A B S T R A C T

Keywords: Peroxymonosulfate N-doped graphitic carbon Magnetic activator Nonradical degradation Core–shell structure

N-doped carbonaceous materials are highly promising for efficient removal of microorganic contaminants from aqueous solutions by simultaneously serving as adsorbents and catalysts. However, their recollection from aqueous solutions for repeated use is challenging. Here, we designed a magnetic N-doped graphitic carbon core–shell structure (MN-GCCS) for oxidative degradation of various microorganic contaminants via the activation of environmentally friendly peroxymonosulfate (PMS). The magnetic Co3O4 core ensured the easily recollection of MN-GCCS via magnetic separation. Meanwhile, the graphitic carbon ideally prevented the Co3O4 from leaching. MN-GCCS showed great reactivity for pollutant degradation at a wide range of pH values and was particularly active under circumneutral conditions (pH 6–8). MN-GCCS outperformed the commonly known activators of PMS; Under identical conditions, near 100% of atrazine was removed with MN-GCCS, while only 36%, 21%, and 14% of the atrazine were removed with multi-walled carbon nanotubes, CuFe2O4, and Co3O4, respectively. In contrast to the known radical processes, the degradation of contaminants by MN-GCCS–PMS was not mediated by radicals. Instead, a mechanism involving a transition oxidation state of MN-GCCS as the major oxidizing intermediate was proposed. The results from this study suggest a novel and highly efficient nonradical process for oxidative degradation of organic contaminants.

⁎

Corresponding author. E-mail addresses: [email protected] (Y. Feng), [email protected] (K. Shih).

https://doi.org/10.1016/j.cej.2019.123724 Received 2 October 2019; Received in revised form 3 December 2019; Accepted 5 December 2019 1385-8947/ © 2019 Elsevier B.V. All rights reserved.

Please cite this article as: Yong Feng, et al., Chemical Engineering Journal, https://doi.org/10.1016/j.cej.2019.123724

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1. Introduction

2. Materials and methods

The presence of various microorganic contaminants in water bodies poses a great threat for the safety of human being and other creatures [1,2]. In the past decade, remediation technologies that rely on highly oxidizing active species, such as hydroxyl radicals (·OH; 0 E0 = 1.8–2.7 V) [3] and sulfate radicals (SO∙− 4 ; E = 2.5–3.1 V) [4], have shown great potential for the efficient degradation, even mineralization, of these contaminants [5,6]. However, most of the degradation studies were carried out with the concentration of target contaminants far beyond their actual environmental levels. The practical application of these radical-based technologies, particularly for the treatment of microorganic pollutants, is challenging because of the ubiquitous presence of radical scavengers, such as natural organic matters and halide ions [7,8]. It has been recognized that developing treatment processes that own a high selectivity toward target pollutants is an effective approach to improve degradation performance and meanwhile achieve a decent utilization efficiency of oxidants. Recently, reports show that peroxymonosulfate (PMS, HSO− 5 ), one kind of persulfates that is widely used to generate SO∙− 4 [9–11], can directly oxidize a wide range of microorganic contaminants, such as arsenic, sulphonamide antibiotics, and fluoroquinolone antibiotics, under elevated pH values (7–10) [12–15]. Such an oxidation technology is promising for practical water remediation, because PMS, a 2− strong oxidant (E0 (HSO− 5 /SO4 ) = 1.75 V) [16,17], is usually more ∙− selective than both SO4 and %OH. In addition, unactivated PMS does not require a solid catalyst, which is energy efficient and free from diffusion or adsorption control. However, few studies have investigated the practical application of unactivated PMS and its aqueous chemistry (e.g., the influence of common anions and effects of pH) is still unclear. In addition, this technology almost has no reactivity under acidic conditions (pH < 4.0). During the past few years, N-doped carbonaceous materials, such as N-doped carbon nanotubes and N-doped reduced graphene oxides, have been investigated for organic contaminant degradation via the activation of persulfates [18]. The incorporation of N atoms could disrupt the electronic and spin culture of the sp2-hybridized configuration and thus break the chemical inertness of these carbonaceous materials to improve their reactivity [18]. For the activation of PMS, the doping of N atoms could drive PMS activation from a radical-mediated process to a nonradical oxidation system [19]. Because radicals are not involved, common anions, such as Cl−, H2 PO−4 , and HCO−3 , do not influence the performance of nonradical degradation [20]. As N-doped carbonaceous materials are stable under strongly acidic conditions, this material-initiated oxidation can be complementary with the unactivated PMS to produce synergistic results under a wide range of pH values. Additionally, the adsorption effect of N-doped carbonaceous materials toward microorganic pollutants can greatly overcome the diffusion problem usually encountered with metal-based activators [21–23]. However, unlike some metallic activators that can be magnetically recycled, the separation of these carbonaceous materials from aqueous solutions is currently a challenge. In this study, we designed a novel magnetic carbonaceous material that consisted of magnetic N-doped graphitic carbon core–shell structures (MN-GCCSs) for microorganic pollutant degradation. The core in MN-GCCSs, composed of magnetic cobalt oxides, can ensure the composite to be easily separated from aqueous solutions via magnetic separation. The material prepared was fully characterized, and its reactivity was compared with commonly known activators with a series of microorganic compounds (properties shown in Table S1) as target contaminants. On the basis of the reactivity, quenching experiments, and the characterization of catalytic sites, a nonradical mechanism for the degradation of contaminants and activation of PMS on the surface of MN-GCCSs was proposed.

2.1. Chemicals and materials PMS in form of Oxone (KHSO5 · 0.5KHSO4 · 0.5K2SO4), cobalt(II) chloride hexahydrate (98%), nitrobenzene (≥99.0%), sulfadiazine (≥99.0%), bisphenol A (≥99%), potassium iodide (≥99.0%), and 2propanol (≥99%) were purchased from Sigma-Aldrich (St. Louis, MO, USA). The content of PMS in the Oxone was quantified prior experiment using a standard sodium thiosulfate solution. A full list of chemicals used in this study is shown in Note S1. 2.2. Material synthesis and characterization MN-GCCSs were prepared using a simple approach involving thermal treatment of cobalt(II) chloride and dicyandiamide. In a typical procedure, cobalt(II) chloride was mixed with dicyandiamide in an agate mortar. The resulting powders were placed in an alumina boat, heated to 500 °C at a ramp of 5 °C min−1 in a tube furnace under N2 atmosphere, and held for 2 h. Then, the temperature was further increased to 700 °C at 5 °C min−1 and held for another 2 h. After cooling to room temperature at a ramp of 5 °C min−1, the black powders produced were washed with 0.5 M H2SO4 for 24 h to remove surface metallic impurities. The MN-GCCSs obtained were collected by magnetic separation, washed with ultrapure water for several times, and dried in an electronic furnace at 50 °C for 24 h. The major composition and crystallinity of the synthesized materials and commercial MWCNTs were investigated using X-ray diffraction (XRD). The XRD pattern was recorded on a Bruker D8 advanced X-ray diffractometer (Karlsruhe, Germany), in which Cu Kα1,2 was used as the source of X-ray irradiation at 40 kV and 40 mA. Fourier transform-infrared (FT-IR) spectroscopic analysis was carried out on a PerkinElmer Spectrum Two FT-IR spectrometer (Waltham, MA, USA). The morphology was investigated with a FEG scanning electron microscope (SEM; Hitachi S-4800) and a scanning transmission electron microscope (TEM; FEI Tecnai G2 20 S-TWIN). X-ray photoelectron spectroscopy (XPS) was recorded on an ECSALAB 250Xi X-ray photoelectron spectrometer (Thermo Scientific, Waltham, MA, USA) with monochromatic Al Kα irradiation. The Brunauer–Emmett–Teller (BET) specific surface area was measured via nitrogen adsorption at 77 K with a TriStar II Plus instrument (Micromeritics, Norcross, GA, USA). The magnetic property of the MN-GCCSs sample was evaluated using a Quantum Design MPMS XL7 magnetometer (San Diego, CA, USA). The potential leaching of cobalt ions from MN-GCCSs during the catalytic degradation was evaluated using inductively coupled plasma-mass spectrometry (ICPMS). The Raman spectrum was record on a LabRAM HR800 laser confocal micro-Raman spectrometer (Horiba Jobin Yvon). The elemental compositions were measured using an elemental analyser (Vario EL Cube, Elementar Analysensysteme GmbH, Germany). 2.3. Experimental procedures Unless otherwise specified, all experiments were carried out at room temperature (25 ± 1 °C). Stock solutions of PMS (500 mM) and organic contaminants were prepared by dissolving exact amounts of Oxone and contaminant powders, respectively, into ultrapure water. In a typical degradation test, 200 mL of the diluted bisphenol A solution (4 μM) was added to a 250-mL glass reactor and the stock solution of PMS was spiked into the solution to reach a final concentration of 1 mM. The pH value of the resulting solution was adjusted using diluted a NaOH or H2SO4 solution. The glass reactor was then placed on an orbital shaker at 250 rpm, followed by the addition of activators to initiate degradation. At predetermined time intervals, around 1 mL aliquots of aqueous sample were withdrawn, filtered with polytetrafluoroethylene (PTFE) membranes, and transferred to 2-mL autosampler glass vials for UPLC analysis. Methanol (100 μL), a strong 2

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of N. The elemental analysis revealed that the content of N was around 2.6% (Table S2). The Co element presented at valent states of both 3+ and 4+ (Fig. 1b). The Co 2p3/2 and Co 2p1/2 main peaks were located at 779.6 and 795.6 eV, respectively, and no satellite peak at around 790 eV was observed. These observations suggest that the cobalt component in MN-GCCSs was Co3O4 [25]. The N was in forms of pyridinic N, pyrrolic N, graphitic N, and quaternary N with atomic percentages of 5.0%, 32.9%, 41.6%, and 20.5% (Fig. 1c), respectively. The FT-IR spectrum shows the presence of C]N, C]C, C–O, C–N, and C–H chemical bonds in MN-GCCSs (Fig. 1d). No peak is observed in the range of 1700–1750 cm−1, which suggests that MN-GCCSs did not contain C]O groups at a quantifiable level [26]. The Raman spectrum of MN-GCCSs shows the characteristic D band (~1360 cm−1), G band (~1600 cm−1), and 2D band (~2708 cm−1) (Fig. 1e) [27]. The D band is a signature of structural defects on the graphitic planes. The ID/IG ratio was measured to be 0.96, which suggests that there were some structural defects on MN-GCCSs. The magnetic property of MN-GCCSs was revealed by the measured saturation moment per unit mass of around 15.5 emu g−1. As revealed by the SEM image, MN-GCCSs had a flake-like morphology (Fig. S2). The metal oxide in the material was enclosed by the graphitic carbon (Fig. 1f), the layer number of which was around 6 (inset in Fig. 1f). Fig. S3 shows the SEM and TEM images of multi-wall carbon nanotubes (MWCNTs); their diameter was in the range of 20 to 30 nm. The energy-dispersive X-ray spectroscopy spectrum reveals that cobalt and copper were present in MWCNTs as metallic impurities (Fig. S4). The specific surface areas of MN-GCCSs, MWCNTs, CuFe2O4, Co3O4 were measured to be 107.5, 185.6, 22.3, and 27.9 m2 g−1 (Fig. S5 and Table S3), respectively.

radical scavenger, was added immediately to the vials to prevent the target contaminant from further degradation. To investigate the role of radicals, different alcohols were added as scavengers to the working solution of bisphenol A during the diluting procedure. To evaluate the stability and reutilization capability of MN-GCCSs, the solids used were recollected by magnetic separation and the supernatant was discarded. The next cycle was initiated by the addition of bisphenol A solution and PMS to the glass reactor that contained the recollected MN-GCCSs. Other conditions were the same as that used in the first catalytic run. 2.4. Chemical analysis The concentrations of bisphenol A, atrazine, sulfadiazine, and nitrobenzene were measured using a Waters Acquity ultra performance liquid chromatography (UPLC) with a photodiode array (PDA) detector (Waters, Milford, MA, USA). The elution of these compounds was carried out on a Waters BEH C18 column (50 mm × 2.1 mm, 1.7 μm) at 50 °C. An injection volume of 10 µL and a flow rate of 0.4 mL min−1 were used throughout the analysis. The detection limits for bisphenol A, atrazine, sulfadiazine, and nitrobenzene were measured to be 0.30, 0.22, 0.11, 0.32 μM, respectively. Detailed analytical parameters can be found in our previous report [24]. 3. Results and discussion 3.1. Properties of MN-GCCSs and other activators Graphitic carbon and cobalt were identified by the XRD pattern (Fig. 1a). The bands at around 26.5° and 44.5° were assigned to the (0 0 2) plane of a graphitic carbon material (JCPDS no. 41–1487) and (1 1 1) plane of cobalt (JCPDS no. 89–7093), respectively. The XPS survey spectrum of MN-GCCSs shows that this material composed of C, N, O, and Co elements (Fig. S1), which confirmed the successful doping

3.2. Reactivity of MN-GCCSs for PMS activation The reactivity of MN-GCCSs was characterized by investigating the degradation of representative microorganic contaminants including bisphenol A, atrazine, and sulfadiazine. Under highly acidic conditions

Fig. 1. (a) XRD pattern and (b, c) high-resolution XPS spectra of MN-GCCS sample, (d) FT-IR and (e) Raman spectra of MN-GCCSs, and (f, inset in f) TEM images of MN-GCCSs. 3

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Fig. 2. (a) Degradation of organic contaminants by MN-GCCSs–PMS and (b) comparison of the reactivity of different materials toward PMS activation. Conditions: [PMS] = 1 mM, [MN-

GCCSs] = [CuFe2O4] = [Co3O4] = [MWCNTs] = 25 mg L−1, [bisphenol A] = [atrazine] = [sulfadiazine] = 4 µM, and pH = (a) 3.0 or (b) 6.8 (10 mM phosphate buffer).

combination of MN-GCCSs with PMS. Nitrobenzene is hard to be oxi6 −1 −1 dized by SO∙− s ) [33,34]. The readily degradation of 4 (< 10 M nitrobenzene excluded the involvement of free/surface-bound SO∙− 4 as the major active species. The near complete failure of these alcohols to inhibit the degradation suggests that neither %OH nor SO∙− 4 contributed to the degradation of contaminants by MN-GCCSs–PMS. It has been reported that radicals (%OH, SO∙− 4 ) are the major active species produced during the activation of PMS by reduced graphene oxides, and the electron rich ketonic groups (C]O) plays a dominant role during such activation [35]. The ketonic groups first form hydrogen bonds with PMS (C]O–H-O-SO3), which then transfer electrons to PMS to generate SO∙− 4 . The absence of radicals in MN-GCCSs–PMS oxidation suggests that the electron-rich groups (such as ketonic groups, if any) and sp2 hybridized carbon framework played a negligible role during the activation. Meanwhile, it is known that SO∙− 4 is the active species generated when PMS is activated by Co2+ or cobalt oxides [5,36]. The nonradical degradation of bisphenol A suggests that the cobalt oxides in MN-GCCSs did not contribute to the activation of PMS, which was probably because that the graphitic layer was too thick to act as an electron tunnel for electron transfer from the cobalt oxides to PMS. This conclusion is in a good agreement with previous findings that the electrons from the core can hardly penetrate multiple shell layers of graphene [18]. The ICP-MS analysis of the reaction solution after catalysis showed that the cobalt ion leached from MN-GCCSs was lower than 40 ng L−1. The catalytic contribution from such a trace level of cobalt ions was negligible. In addition to radicals, singlet oxygen (1O2) is also a common species produced by the activated PMS [23,37]. To study the role of 1O2, the effect of furfuryl alcohol (FFA) on the degradation of bisphenol A by MN-GCCSs-PMS was explored. FFA reacts rapidly with 1O2 (1.2 × 108 M−1 s−1) [38] and has thus been widely used to identify the role of 1O2 [39]. Meanwhile, very recent results from Yang et al. [40] show that this alcohol could also react with PMS, although the reaction highly depends on the existing forms of both reactants. For such a cause, relatively low levels of FFA were studied. The oxidation of FFA by PMS can be expressed by a second-order law (Eq. (1)):

(~pH 3.0), PMS alone had almost no oxidation of these compounds (Fig. S6). However, the adsorption of MN-GCCSs toward organic pollutants was observed in the absence of PMS; around 15.7, 11.4, and 19.4% of the bisphenol A, atrazine, and sulfadiazine, respectively, were removed after contacting with MN-GCCSs for only 15 min (Fig. S7). When MN-GCCSs was co-present with PMS, rapid degradation of target contaminants occurred (Fig. 2a); approximately 90%, 99%, and 100% of the bisphenol A, atrazine, and sulfadiazine, respectively, were removed after reaction for 15 min in the presence of 1 mM PMS and 25 mg L−1 MN-GCCSs. The efficient removal of target contaminants indicates the great reactivity of MN-GCCSs toward PMS activation. To further evaluate the reactivity of MN-GCCSs for PMS activation, we compared this material with other commonly known efficient activators including CuFe2O4, Co3O4, and MWCNTs. During such evaluation, atrazine was selected as a target contaminant. Under identical conditions, the MN-GCCSs had significantly higher reactivity than the other three materials. More than 99% of the atrazine was removed by MN-GCCSs–PMS after reaction for 1 min, whereas only around 36%, 21%, and 14% of the atrazine were removed by MWCNTs-, CuFe2O4-, and Co3O4-activated PMS, respectively (Fig. 2b). Interestingly, MWCNTs showed higher reactivity than Co3O4, which was probably due to the synergy between adsorption and activation degradation in MWCNTs-PMS. The superior of some carbonaceous materials, such as reduced graphene oxide and CNT, over Co3O4 in the activation of PMS was also previously observed by other researchers [28,29]. 3.3. Reactive species for contaminant degradation To understand the activation of PMS by MN-GCCSs, it is prerequisite to study the generated dominant active species. To examine the role of radicals, we conducted classical scavenging experiments using different alcohols. Alcohols such as methanol, ethanol, and propanol react rapidly with radicals, such as ·OH (kmethanol = 9.7 × 108 M−1 s−1, kethanol = 1.9 × 109 M−1 s−1, kpropanol = 1.9 × 109 M−1 s−1) [3] and SO∙− (kmethanol = (2.5 ± 0.4) × 107 M−1 s−1, 4 kethanol = (7.7 ± 2.2) × 107 M−1 s−1, 7 −1 −1 kpropanol = (8.5 ± 3.0) × 10 M s ) [30]. Therefore, the effects of these alcohols on the oxidative performance of MN-GCCSs–PMS were studied. The results show that high levels of alcohols had only a negligible influence, as revealed by the near constant degradation of bisphenol A regardless the alcohols were present or not (Fig. 3a). This phenomenon was also observed when the concentration of methanol was further increased to 10 M (Fig. 3b). Although radicals react rapidly with bisphenol A (k∙OH=(7.2 ± 0.34 ) × 109 M−1 s−1 [31], k SO∙− =(1.37 ± 0.15) × 109 M−1 s−1 [32]), such high levels of alcohols 4 were sufficient to quench its radical-induced degradation. Furthermore, degradation experiments with nitrobenzene as a target pollutant showed that this compound could also be efficiently degraded by the

−

d[FFA] = kFFA,PMS [FFA][PMS] dt

(1)

when [FFA] = 0.1 mM, [PMS] ˃˃ [FFA]. Thus, Eq. (1) can be represented as:

−

d[FFA] = k obs [FFA] dt

(2)

where k obs (kFFA,PMS [PMS]) is the pseudo first-order rate constant between PMS and FFA. After integration, Eq. 2 can be represented as:

[FFA] = [FFA]0 e−k obs t 4

(3)

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Fig. 3. (a, b) Effects of different alcohols on the degradation of bisphenol A by MN-GCCSs–PMS, (c) degradation of nitrobenzene by MN-GCCSs–PMS, and effects of furfuryl alcohol on the degradation of bisphenol A by MN-GCCSs–PMS. Conditions: [PMS] = 1 mM, [MN-GCCSs] = 25 mg L−1, [bisphenol A] = 4 µM, (b) [methanol] = 10 M, and pH 3.0.

to the oxidation by unactivated PMS. To better quantify the contribution of activated PMS, we investigated the degradation of different contaminants by PMS alone under different pH conditions, which is usually neglected in the investigation of activated PMS. Under acidic conditions (~pH 3.0), only slight degradation of bisphenol A and sulfadiazine occurred (Fig. S6). When the pH of the solution was in the range of 4.5–10.0, all the three compounds including bisphenol A, atrazine, and sulfadiazine were observed to be obviously degraded (Fig. S8 to Fig. S10). The degradation of these compounds followed pseudo-first-order kinetics. The pseudofirst-order rate constants calculated demonstrate that unactivated PMS had the highest reactivity of degrading these compounds at around pH 6.8 (Fig. 4b and Fig. S11). Under basic conditions PMS undergoes self-decomposition to generate 1O2 as the active species (Eq. (5), k = 0.013 ± 0.0003 M−1 s−1) [40,41,43]. The stoichiometry of this reaction was recently verified by Yang et al. [40] Previous studies have reported the degradation of organic contaminants by PMS under basic conditions [12,15] and some studied proposed that 1O2 plays an important role during such degradation [14]. The pKa,2 of PMS is around 9.4, and thus the generation of 1O2 has the highest kinetics at around pH 9.4. Meanwhile, atrazine

With Eq. (3), the half-life of FFA can be calculated as:

t1/2 = ln2/ k obs

(4)

The pKa,2 and pKa of PMS and FFA are around 9.4 [41] and 9.55 [42]; Under the acidic conditions investigated, FFA presented at a neutral form and its second-order rate constant with PMS (kFFA, PMS) was reported to be 0.059 ± 0.004 M−1 s−1 [40]. When FFA was present at a concentration of 0.1 mM, its half-life was calculated to be more than 32 h. The half of the initial concentration of FFA was still more than 10 times greater than the level of bisphenol A (4 µM). However, no inhibitory effects from the FFA were observed. Instead, the FFA displayed a slight promoting effect. The absence of any adverse effect from FFA suggests that 1O2, if any, did not contributed to the degradation of bisphenol A. The pH value is a very important factor in aquatic chemistry and thus the reactivity of MN-GCCSs under different pH values was evaluated. When the pH value was increased from acidic condition (pH 3.0) to near neutral and basic conditions, the degradation of target compound atrazine was significantly increased from around 40% to more than 99% after reaction for only 2 min (Fig. 4a). The promoted degradation of atrazine under elevated pH conditions was in part ascribed

Fig. 4. Reactivity of MN-GCCSs toward PMS activation under different pH conditions. Conditions: [PMS] = 1 mM, [MN-GCCSs] = 25 mg L−1, [atrazine] = 4 µM, [acetate buffer] = 5 mM (pH 4.5), [phosphate buffer] = 5 mM (pH 6.8, 8.0), and [boric buffer] = 5 mM (pH 10). 5

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carbon shells, which could lower the local work function of the outer carbon surface and facilitate its interaction with PMS and organic substances [47]. On the one hand, the pyrrolic N present in carbonaceous materials was previously found to be the active site for adsorbing organic contaminants [23], which explained the adsorption removal of pollutants by MN-GCCSs. On the other hand, the graphitic N in MNGCCSs was sp2-hydridized and possessed lone pair electrons, which formed delocalized conjugated π systems with the sp2-hydridized graphitic carbon frameworks [19]. Therefore, the graphitic N had high electronegativity, which could abstract electrons from the adjacent carbon atoms. In contrast to previous non-radical processes in which the presence of an organic substrate has a positive effect on the decomposition of either PDS or PMS [48], more rapid decomposition of PMS was recorded in the MN-GCCS suspensions in the absence of a contaminant (Fig. S13), which indicates that the decomposition of PMS did not rely on the electron transfer from the organic substrate. On the basis of the above investigations, we propose a nonradical two-step mechanism for pollutant degradation on MN-GCCSs. First, the electron-rich graphitic N interacted with PMS to form a transition oxidation state by acting as an electron donor. Second, the transition state formed then oxidized the contaminants that were adsorbed onto the surface of MN-GCCSs. As both activation and degradation were surface reactions, competition between the adsorption of PMS and contaminants onto MN-GCCSs was expected to exist, which explains the inhibiting effect of bisphenol A on the decomposition of PMS (Fig. S13). Meanwhile, electron transfer probably occurred in the structure of the transition state, resulting in the generation of surface oxygen-containing groups. As revealed by the XPS data, the atomic ratio between O and C increased from 0.04 in the pristine MN-GCCSs to 0.18 in the MNGCCSs after the 10th cycle (Fig. S14). A sp2 carbon framework with increased oxygen species (high oxidation state) was previously found to be unfavorable to undergo electron transfer to activate PMS [49], which was probably in part responsible for the deactivation observed in the 10th activation cycle (Fig. 5a). However, such oxygen groups can be easily removed/reduced by a thermal treatment [50,51].

has a pKa value of around 1.6 (Table S1); this compound existed primarily as neutral molecules throughout the pH range tested. Therefore, if the 1O2 generated via Eq. (5) was responsible for the degradation of atrazine, the best result was expected to achieve under basic conditions (pH 8 or 10). However, as described above, the best degradation of atrazine and the other compounds was observed at pH 6.8. This discrepancy was probably caused by the relatively high levels of scavengers used in previous reports to identify the role of active species [14]. In addition to act as a scavenger for the target species, the compound added also invalidly consume PMS significantly. For example, sodium azide, a widely used scavenger of 1O2 (4.5 × 108 M−1 s−1), reacts with PMS under acidic and neutral conditions (k = 0.66 M−1 s−1) [44]. In addition, the 1O2 produced can be rapidly quenched by water (2.5 × 105 s−1) [45], which is usually failed to consider when evaluating the contribution of 1O2. The mechanism underling the optimal condition of pH 6.8 is still unclear, but is probably associated with several factors such as the pKa values of reactants − 2− (Table S1) and the quite different reactivity of HSO− 5 and SO5 ; HSO5 0 2− (E = 1.75 V) is more electrophilic and oxidative than SO5 (E0 = 1.22 V) [16]. The presenting form of organic contaminants has a significant effect on their interaction with PMS. Recent results from Zhou et al. [13] show that the anionic forms of ciprofloxacin and enrofloxacin react with PMS more than 19 and 6 times, respectively, faster than the neutral forms. 2− 2− + HSO− 5 + SO5 → 2SO4 + 1O2 + H

(5)

3.4. Stability and functioning mechanisms of MN-GCCSs The reactivity of MN-GCCSs in the multiple catalytic cycles was investigated. No obvious deactivation occurred during the beginning 7 cycles (Fig. 5a), which suggests that MN-GCCSs had high stability. However, a sharp drop in the degradation of bisphenol A was noticed in the continuous operation; the degradation rate was decreased from 94.0% to 25.5% and 7.8% in the 8th and 9th cycles, respectively. The significant decrease suggests the deactivation of activators. Previous studies by Lee et al. [46] show that the deactivation of carbonaceous catalysts is partially caused by accumulation of target pollutants, degradation products, and persulfates. We therefore tried to regenerate the MN-GCCSs after the 9th cycle by washing the material using methanol and ultrapure water under sonication. As shown, most of the catalytic reactivity of MN-GCCs could be recovered; the degradation rate of bisphenol A increased from 7.8% to 75.3% in the 10th cycle. The significant recovery of the catalytic reactivity suggests that the deactivation of the material was probably mainly caused by the accumulation of reactants and degradation products, which prevented the exposure of surface activation sites. However, compared with the performance of the fresh material, the degradation rate in the 10th was still more than 20% lower. To explore the cause, we studied the morphology of the used MN-GCCSs. The TEM image shows that the cobalt oxides were still well entrapped by the N-doped graphitic carbon. However, compared with the pristine MN-GCCSs (Fig. S12), aggregation and restacking of the material were observed after the catalytic process (Fig. 5b), which led to an increase in the thickness of the graphitic shell layer and a decrease in the numbers of exposed activation sites. The XRD pattern of the used MN-GCCSs was recorded and the results show that the signal of cobalt was relatively decreased (Fig. 5c). The saturation magnetic moment of the MN-GCCSs after 10th run was measured to be 13.1 emu g−1, slightly lower than the value of the fresh material (15.4 emu g−1). All of these observations were probably caused by the restacking of the graphitic shell layer. Although the saturation magnetic moment of MN-GCCSs was slightly decreased after reaction, such level of magnetism can still ensure the magnetic recycle of the composite (inset, Fig. 5d). In addition to provide magnetic property, the cobalt oxides in MNGCCSs were expected to interact with the inner sphere of graphitic

4. Conclusions This study investigated the activation of PMS by MN-GCCSs for the degradation of several microorganic contaminants. Some conclusions can be drawn: (1) MN-GCCSs had great reactivity for pollutant degradation at a wide range of pH values and was particularly active under circumneutral conditions (pH 6–8). (2) MN-GCCSs were more effective than the commonly known activators of PMS, such as multi-walled carbon nanotubes, CuFe2O4, and Co3O4. (3) In contrast to previous reports, neither radical nor 1O2 was the active species generated by MN-GCCSs–PMS. Instead, a transition oxidation state of MN-GCCSs was probably the major oxidizing intermediate that was responsible for the degradation of contaminants. (4) Contaminants including bisphenol A, atrazine, and sulfadiazine can be directly oxidized by unactivated PMS, and the best performance was achieved at circumneutral conditions. (5) MN-GCCSs showed high reuse capability with around 94% removal rate of bisphenol A after the 7th catalytic cycle. (6) The graphitic shell in MN-GCCSs well prevented the cobalt from leaching, and the Co3O4 core ensured the easily magnetic recollection of MN-GCCSs. Acknowledgements This study was funded by the Research Grants Council of Hong Kong (Projects 106180082, C7044-14G, and T21-711/16R) and the start-up 6

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Fig. 5. (a) Reactivity of MN-GCCSs in multiple cycles, (b) TEM image of the MN-GCCSs after 10th run, (c) XRD patterns of fresh and used MN-GCCSs, and (d) magnetic hysteresis loops of fresh and used MN-GCCSs. Inset in (d) shows the MN-GCCSs-PMS suspensions before and after magnetic separation. Conditions: [PMS] = 1 mM, [MN-GCCSs] = 25 mg L–1, and [bisphenol A] = 4 µM.

fund from South China Normal University (8S0597).

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