Biochar-supported nanoscale zero-valent iron as an efficient catalyst for organic degradation in groundwater

Journal of Hazardous Materials 383 (2020) 121240

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Biochar-supported nanoscale zero-valent iron as an efficient catalyst for organic degradation in groundwater

T

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Zhe Lia,b,c, Yuqing Sunc, Yang Yanga,b, Yitong Hana,b, Tongshuai Wanga,b, Jiawei Chena,b, , ⁎⁎ Daniel C.W. Tsangc, a

State Key Laboratory of Biogeology and Environmental Geology, China University of Geosciences, Beijing, 100083, PR China School of Earth Sciences and Resources, China University of Geosciences, Beijing, 100083, PR China c Department of Civil and Environmental Engineering, The Hong Kong Polytechnic University, Hung Hom Kowloon, Hong Kong, China b

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

A R T I C LE I N FO

A B S T R A C T

Editor: Danmeng Shuai

High-efficiency and cost-effective catalysts are critical to completely mineralization of organic contaminants for in-situ groundwater remediation via advanced oxidation processes (AOPs). The engineered biochar is a promising method for waste biomass utilization and sustainable remediation. This study engineers maize stalk (S)and maize cob (C)-derived biochars (i.e., SB300, SB600, CB300, and CB600, respectively) with oxygen-containing functional groups as a carbon-based support for nanoscale zero-valent iron (nZVI). Morphological and physiochemical characterization showed that nZVI could be impregnated within the framework of the synthesized Fe-CB600 composite, which exhibited the largest surface area, pore volume, iron loading capacity, and Fe0 proportion. Superior degradation efficiency (100% removal in 20 min) of trichloroethylene (TCE, 0.1 mM) and fast pseudo-first-order kinetics (kobs =22.0 h−1) were achieved via peroxymonosulfate (PMS, 5 mM) activation by the Fe-CB600 (1 g L−1) under groundwater condition (bicarbonate buffer solution at pH = 8.2). Superoxide radical and singlet oxygen mediated by Fe0 and oxygen-containing group (i.e., C]O) were demonstrated as the major reactive oxygen species (ROSs) responsible for TCE dechlorination. The effectiveness and mechanism of the Fe/C composites for rectifying organic-contaminated groundwater were depicted in this study.

Keywords: Biomass waste valorization Engineered biochar nZVI-carbon composites Sustainable waste management Reactive oxygen species Sustainable/green remediation

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Corresponding author. Corresponding author at: State Key Laboratory of Biogeology and Environmental Geology, China University of Geosciences, Beijing, 100083, PR China. E-mail addresses: [email protected] (J. Chen), [email protected] (D.C.W. Tsang).

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https://doi.org/10.1016/j.jhazmat.2019.121240 Received 5 July 2019; Received in revised form 12 September 2019; Accepted 14 September 2019 Available online 17 September 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.

Journal of Hazardous Materials 383 (2020) 121240

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1. Introduction

et al., 2005; Liang et al., 2018). However, excessive structural defects may compromise the structural integrity/durability and mechanical strength of the carbon-based catalysts. Besides, the initial contact between organic pollutants and carbon matrix through adsorption has been demonstrated as a key step in catalytic organic degradation. The task in designing a new generation of biochar supported materials for green environmental remediation is challenging. The interdependent relationships between the physicochemical structural properties, adsorptive/catalytic performance, and crosslinking mechanisms remain unclear. This paper envisions unraveling some of the current research limitations in the synthesis and application of the Fe/C composites for sustainable environmental remediation. In this study, waste maize straws and maize cobs were used as the raw materials for the designed biochar with layer structure for supporting nZVI via liquid phase reduction. The characteristics of resultant biochar-supported nZVI (Fe/C) and synergistic effect on their efficiency of PMS activation were scrutinized in terms of the removal of trichloroethylene (TCE), a common organic contaminant in groundwater under field-relevant conditions. Our study intends to develop an efficient and sustainable carbon-based catalyst for organic degradation and realize the following specific objectives: (1) fabricate various Fe/C composites via different pyrolysis temperatures (i.e., 300 and 600 °C) and biochar precursors (i.e., corn straw and corncob), and examine the organic degradation efficiency to elucidate possible synergistic interactions between nZVI and biochar; (2) determine the effects of PMS dosages, Fe-to-biochar impregnation ratios, initial TCE concentrations, reaction temperature, and coexisting inorganic anions (Cl−, NO3−, PO43−, and HPO42−) on the effectiveness of the Fe/C composites; and (3) identify the relative contribution of individual ROSs to the PMS activation and catalytic degradation of TCE via scavenging text and electron spin resonance (ESR) analysis. The results of this paper could make improvements to the design of high performance engineering biochar composites for groundwater remediation.

Advanced oxidation processes (AOPs) have been widely used because of its enormous potential capability in the completely mineralization of varies organic contaminants by the generation of highly reactive oxygen species (ROSs), such as hydroxyl radical (%OH), sulfate radical (SO4%−), superoxide radical (O2%−), and singlet oxygen (1O2) (Gogate and Pandit, 2004). Among these radicals, SO4%− has higher selectivity in decomposing target specific functional groups in organic compounds compared with %OH, owing to its higher redox potential (2.5–3.1 V vs. 1.8–2.7 V), wide pH applicability, and adaption to flexible conditions (Sun et al., 2019a). In general, SO4%− can be generated by activated persulfate (PS) or peroxymonosulfate (PMS) through transition metals, ultraviolet, heat, and ultrasound. Transition metals (e.g., Fe, Co, and Mn) were chosen to active PMS as this process require lower activation energy and highly efficient for activation (Oh et al., 2016; Zhou et al., 2019). Iron-based materials as heterogeneous catalyst have been widely used because iron presents advantages of being environmentally friendly, low-cost, non-toxic, and highly accessible as the second most abundant metallic element of the earth crust (Fu et al., 2014). In particular, nZVI shows higher reaction efficiency for activating PMS than the iron oxide and the larger ZVI owing to its characteristics of large surface area to particle size ratio (Barzegar et al., 2018; Jaafarzadeh et al., 2017). The magnetic property of nZVI makes it easier to separate and recycle from the aqueous solutions under external magnetic (Lei et al., 2018). However, conventional nZVI’ s intrinsic chemical and physical properties would hind its advantages. It would cause agglomeration and lead to the formation of larger particles with a significant drop in reactivity during the reaction progress (Reddy et al., 2016; Sun et al., 2017, 2018). Recently, carbon based materials such as activated carbon (AC), graphene oxide (GO), and carbon nanotubes (CNTs) have attracted much attention due to their high surface area and good porosity for catalyst dispersion (Sun et al., 2012; Xi et al., 2014). The utilization of renewable agricultural residues (i.e., food supply chain wastes) for sustainable waste management and environmental remediation has gained great attention. Statistics from Chinese Academy of Agriculture Science in 2005 indicated that large numbers of byproduct from corn production (maize straw and maize cob) accounted for 34.2% of 718 million tons of crop wastes (Yang et al., 2018). Biochar prepared from maize straw and maize cob with a designed pyrolysis under limited oxygen conditions will have high specific area, oxygen-containing functional groups, and stable structure for pollutants removal (Yang et al., 2019a). To use biochar as the supporting material for nZVI will provide nZVI with a good dispersion and higher efficiency (Sun et al., 2019b). Recent studies suggested that various carbonaceous materials (e.g., AC, CNTs, and GO) are capable of catalyzing PMS for organic degradation via electron transfer through graphitic carbon matrix, which can be even more efficient than some transition metal oxides (Oh et al., 2016; Duan et al., 2016; Yang et al., 2016). Compared with other carbon materials, biochar has more oxygen-containing functional groups on its surface (e.g., eCO, and C] eOOH), which may also be capable of interacting with PMS to generate ROSs (Hussain et al., 2016; Yan et al., 2017). There are still many uncertainties in tailoring biochar-supported nZVI for catalytic organic degradation in the latest literature. Biochar derived from various feedstock via pyrolyzing at different temperatures can lead to diverse physicochemical properties, porous structure, and surface chemistry (Ruan et al., 2019). On one hand, these distinctive characteristics may influence the affinity between carbon matrix and anchored nZVI particles (Yang et al., 2019b). The catalytic activity of biochar is highly dependent on its intrinsic (graphitization degree, surface chemistry, etc.) and extrinsic (specific surface area, morphology, etc.) properties. The structural defects in the carbonaceous matrix can host the active redox-pairs/cycle (e.g., Fe0-Fe2+-Fe3+), thus increasing the density of the active sites for PMS activation (Banerjee

2. Materials and methods 2.1. Chemical and materials Maize straws and maize cobs collected from Hebei Province (China) were selected as the feedstock. In this study, the corn straw and cob, highly abundant agricultural waste with low ash content, were selected as biochar precursors to avoid the potential interferences of other coexisting metal ions. Before use, maize straws and maize cobs were washed by deionized (DI) water for several times and dried at 60 °C for 24 h, crushed into small piece (< 5 mm), and stored in airtight containers before use. The following chemical reagents were of analytical grade and purchased from Sinopharm Chemical Reagent Co., Ltd. (viz. ferrous sulfate heptahydrate (FeSO4·7H2O), sodium borohydride (NaBH4), sodium bicarbonate (NaHCO3), sodium hydroxide (NaOH), sodium chloride (NaCl), sodium nitrate (NaNO3), trisodium phosphate dodecahydrate (Na3PO4·12H2O), disodium hydrogen phosphate dodecahydrate (Na2HPO4·12H2O), sulfuric acid (H2SO4), tert-butyl alcohol (C4H10O, TBA), methanol (CH4O), and ethanol (C2H6O, EtOH)), Aladdin (China) (viz. TCE), and Alfa Aesar (China) (viz. PMS (2KHSO5·KHSO4·K2SO4), L-Histidine (C6H9N3O2), and p-benzoquinone (C6H4O2, PBQ). DI (18.2 MΩ cm−1) used in experiment was produced by our own lab for all solutions. 2.2. Preparation of catalyst Two kinds of biomass, maize straw and maize cob, were selected as the feedstocks of biochar. After cut, washed, and dried at 60 °C for 24 h, biomass was pyrolyzed in muffle furnace at 300 °C and 600 °C for 2 h, respectively. The temperature of biochar synthesis was set at 300 °C and 600 °C, which represent low and high pyrolytic temperatures commonly used for biochar preparation, to study the influence of pyrolysis 2

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temperature on the resultant Fe/C composites. The particle size of maize straw biochar (SB) and maize cob biochar (CB) selected in this study was 0.3 − 0.5 mm and referred to as SB300, SB600, CB300, and CB600, respectively. The preparation of Fe/C composites was exhibited as follows: 1 g biochar (SB300, SB600, CB300, and CB600, respectively) and 4.9643 g FeSO4·7H2O were dissolved and shaken for 12 h in a mixed solvent containing 50 mL DI. After 10-min setting, the precipitates were washed with DI for three times to remove excess Fe. The obtained solids were immerged into 100 mL DI and an additional 40 min of 1000 rmp stirring mixed with 30 mL NaBH4 (2.8896 g NaBH4 dissolved in 0.1% NaOH solution) at 2 drops s−1. The whole process was performed in nitrogen atmosphere. Then the black precipitates were washed with deoxygenated DI and ethanol three times and vacuum dried for 24 h at the temperature of -60 °C. The resulting Fe/C composites were referred to as Fe-SB300, FeSB600, Fe-CB300, and Fe-CB600, respectively. For comparison, the nZVI particles were also synthesized following the same procedures.

2.5. Analytical methods

2.3. Characterization of Fe/C composites

Fig. S1 showed the morphology of four Fe/C samples by SEM analysis. The SEM results showed that Fe/C samples had an obvious sheet structure. The surface of biochar was relatively rough and irregular. It can be observed that nZVI particles synthesized by liquid phase reduction dispersed well on the biochar surface without obvious aggregation. In contrast to the as-prepared nZVI particles that would agglomerate together (Sun et al., 2017), biochar’s porous structure effectively avoid the irreversible aggregation and maintain the large surface area and high reactivity of nZVI for organic degradation. Based on the nitrogen adsorption-desorption isotherms, the Brunauer-Emmett-Teller surface area (SBET) and pore volume of biochar and Fe/C composites were shown in Fig. 1. According to the IUPAC classification, it revealed that the adsorption isotherms of all Fe/C samples were considered characteristic type IV isotherms with a H4 hysteresis loop implying that the existence of micropores and mesoporous in the biochar and Fe/C composites. The majority of pore diameters were distributed in the ranges of 2–10 nm (Table S2). It further confirmed significance of mesopore distribution in these Fe/C composites. According to the porosity analysis, the BET surface area (SBET), total pore volume (TPV), and average pore diameter (Ad) of biochar increased with elevated pyrolysis temperature from 300 °C to 600 °C. Higher pyrolysis temperature would enhance the biomass thermochemical decomposition and improve porosity development (Yang et al., 2019c). Compared with the SB600, it is clear seen that the SBET (8.15 m2 g−1), TPV (0.0523 cm3 g−1), and Ad (2.57 nm) of the FeSB600 were notably decreased due to the impregnation of iron particles, which blocked the micropores of the raw biochar. However, the less porous SB300 structure was less affected and manifested a slightly decreased SBET (9.55 m2 g−1) and enlarged Ad (8.12 nm) of Fe-SB300. The obvious increase in the TPV of the Fe-CB600 (0.205 cm3 g−1) was possibly contributed by the presence of iron particles, which are expected to enhance the organic degradation performance. The iron content of the Fe/C composites ranged from 21.4 to 38.6 wt%, which was lower than the 50.0% Fe mass ratio present in the Fe/biochar upon theoretical value (Fig. 2). This loss of Fe was possibly owing to the post-synthesis washing steps that removed loosely bound and excess Fe. The Fe loading was generally greater on the 600 °Ccomposites as compared with 300 °C-composites. Moreover, the Fe content significantly increased from 22.8% on the Fe-CB300 to 38.6% on the Fe-CB600 owing to the increased porosity of biochar matrix. The disordered structure of the biochar and Fe/C composites was revealed by the Raman spectroscopy (Fig. 3). The D peak at 1358 cm−1 and G peak at 1585 cm−1 correspond to the in-plane vibrations of the disordered sturctures in carbon and graphite (Gangupomu et al., 2016; Gupta et al., 2014; Tavares et al., 2015) respectively. Higher ID/IG ratio suggests that high-temperature pyrolysis and Fe impregnation could induce a more disordered structure with lower graphitization due to more porous and sheet-like morphology (Ouyang et al., 2019).

The concentration of TCE was measured with headspace sampling by gas chromatograph equipped with a flame ionization detector (GCFID). The concentration of chloride due to TCE degradation during the batch experiments was measured by the ionic chromatography (IC, DIONEX ICS-900, Thermo Fisher, USA). The Fe content in the Fe/C composites was determined by an inductively coupled plasma atomic emission spectroscopy (ICP-AES). The ROS generated during the reaction were obtained by electron spin resonance (ESR). Detailed analysis parameters were shown in the Supplementary Information. All experiments were performed in triplicate and the results were the mean with standard error. 3. Results and discussion 3.1. Characteristics of the Fe/C composites

The specific surface area of nZVI and biochar-supported nZVI were measured by N2 adsorption using a Quadrasorb Station 1 analyzer. Morphology and elemental composition were observed by scanning electron microscopy (SEM) with energy-dispersive X-ray spectroscopy (EDS) operated at 20 kV. Raman spectra were recorded under ambient conditions using a confocal micro-Raman spectrometer (Renishwa Invia Fellex, UK) equipped with a Diode-Pumped Solid State (DPSS) laser source emitting at 532 nm. To characterize the functional groups on Fe/ C surface, Fourier transform infrared spectroscopy (FTIR) was used with the scan rage from 4000 to 400 cm−1. X-ray photoelectron spectra (XPS) were obtained using Escalab 250 Xi (Thermo Fisher) with Al Kα radiation. X-ray diffraction (XRD; D8 Advance diffractometer, Bruker) was used to investigate the crystal structure of Fe/C composites using Cu-Ka radiation (l ¼ 1.5406) at 40 mA. Detailed analysis parameters of SEM, XPS and ESR were shown in the Supplementary Information. 2.4. Catalytic degradation of TCE by Fe/C composites The TCE removal experiments were performed at 25 ± 1 (15 or 35) °C under extensive mixing by a dark box shaker (180 rpm). An aliquot of 60 μL TCE stock solution (10 g L−1) in methanol was spiked into bicarbonate buffer solution (pH = 8.2 ± 0.1 with similar water chemistry with groundwater) to obtain the initial concentration of TCE to 0.1 mM. To obtain the degradation kinetics of TCE by the Fe/C, 1.0 g L−1 Fe/C and 5 mM PMS (determined according to preliminary results) were reacted with 100 mL bicarbonate buffer solution containing 0.1 mM TCE. The same condition with nZVI was conducted for comparison in 250 mL serum bottles capped with gas-tight Teflon Mininert valves. Buffer solutions were deoxygenated by purging N2 of ultra-high purity for 5 min before reaction. At different time intervals, a sample of 50 μL headspace gas was withdrawn from the serum bottle and analyzed for TCE concentration. The scavenging experiments were performed to illustrate the contribution of different ROSs (i.e., %OH, SO4%−, O2%−, and 1O2) by using 1 M EtOH (for %OH and SO4%−), 1 M TBA (for %OH), 0.05 M PBQ (for O2%−), and 0.05 M L-Histidine (for 1O2), respectively. After reaction, the spent catalyst was recycled to test its reusability. In order to evaluate the field applicability of the PMS/Fe-CB600 system for groundwater remediation, the time-dependent TCE removal was tested in the simulated groundwater under the same conditions (Table S1). The effects of co-existing inorganic anions (Cl–, NO3–, PO43–, and HPO42–) on the TCE degradation were investigated. The same reactions were conducted in bicarbonate buffer solution containing competing inorganic anions at various concentrations (0, 5, 10, and 20 mM), which were prepared by dissolving NaCl, NaNO3, Na3PO4·12H2O, and Na2HPO4·12H2O, respectively. 3

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Fig. 1. BET spectra of biochars and Fe/C composites.

The XRD patterns of different samples were shown in Fig. S3a indicating that the biochar kept typical biochars structure and the framework of biochars maintained intact. The characteristic peaks at 45°was assigned to α-Fe0 corresponding to the (110), indicating the formation of zero valent iron. The XRD spectra indicated ZVI as the major crystal phase of iron species, which could predominantly contributed to the PMS activation and TCE degradation in this study. FTIR spectra for different Fe/C composites were shown in Fig. S3b. The band at approximately 3400 cm−1 is associated with vibration in hydroxyl groups (OHe). The peak at 1590 cm−1 was assigned to carbonyl/carboxyl C]O stretching vibration. The band at 1103 cm−1 corresponds to the COe stretching vibration modes (Liu et al., 2010). The peaks of Fe–O fall in the range of 800–400 cm−1, comfirming that Fe was attached to the surface by bonding with oxygen-containing groups. The chemical compositions of biochars and Fe/C composites were further studied via XPS analysis (Fig. 4). The binding energies of C 1s at 284.7 eV, 286.0 eV, and 287.2 eV were typically attributed to CeC, CO

Fig. 2. Iron content of Fe/C composites.

Fig. 3. Raman spectra of biochars and Fe/C composites. 4

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Fig. 4. XPS analysis of C 1s and Fe 2p binding states of biochars and Fe/C composites.

3.2. Efficacy of Fe/C composites for PMS activation and catalytic degradation of TCE

and COe] (Huang et al., 2018). As the pyrolysis temperature increased, it showed a decrease in the contents of CeC and CO], whereas the content of COe increased due to the progressive condensation of the aromatic structure at higher temperature, in agreement with previous findings (Shen et al., 2019). In the Fe 2p spectra, the binding energy of characteristic peaks were at 711.50 eV and 710.70 eV for Fe3+ and Fe2+, respectively (Allen et al., 1974; Andersson and Howe, 1989). These results are consistent with the FTIR results. In particular, the peak at 707.00 eV was assigned to Fe0 (Mathieu and Landolt, 1986), which corroborated the impregnation and reduction of Fe2+ to nZVI on the porous surface of biochar. The oxidation of ZVI was inevitable during the aquatic chemical reduction, freeze drying, and storage processes. The previously reported lab-synthesized and commercialized nZVI particles all exhibited a typical core-shell structure, where the core is mainly ZVI and the shell is mixed iron oxides (Bae et al., 2016). As the CeO and CO] fractions of biochars were higher than those of Fe/C composites, the C–O–Fe bonds formed as Fe2+ was impregnated onto the biochar by bidentate chelation with COe, which was then broken by electron transfer during NaBH4 reduction to form nZVI and iron oxides. Interestingly, the highest content of Fe0 (14.1%) was observed for the Fe-CB600, followed by that of the Fe-SB300 (4.1%), Fe-CB300 (2.1%), and Fe-SB600 (1.0%).

The time-dependent removal of TCE by the Fe/C composites was shown in Fig. 5a. In the initial stage (0–1 h), the degradation kinetics exhibited fast rate and then it showed a slower rate for 1–2 h, finally reached an apparent maximum at 3–6 h. TCE degradation efficiency was greater for the 600 °C-Fe/C composites as compared with the 300 °C-Fe/C composites. In addition, TCE removal by the Fe-CB was higher than that of the Fe-SB. The larger SSA (22.0 m2 g−1), TPV (0.205 cm3 g−1), and higher Fe content (38.6%) of Fe-CB-600 may all contribute to more active sites for the superior performance (43.4% removal in 20 min). In contrast, it is interesting to note that the PMS addition significantly enhanced the TCE removal by Fe-SB300 (from 7.70%–80.3% in 20 min), Fe-CB300 (from 12.6%–89.0% in 20 min), Fe-SB600 (from 32.8%–94.4% in 20 min), and Fe-CB600 (from 43.4%–100% in 20 min), respectively. The degradation data was fitted to the pseudo-first-order kinetics model (R2 = 0.92–0.99) and the observed rate constant (kobs) suggested a 37.2–180-fold acceleration of the TCE removal after PMS activation (Fig. 5b). However, control experiments proved that PMS alone could not oxidize TCE in the absence of Fe/C catalysts. It is particularly noteworthy that these catalytic capacities outcompeted the lab-synthesized nZVI for the TCE removal (77.0% in 20 min). This 5

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Fig. 5. Kinetics for TCE removal by the (a) Fe/C, (b) Fe/C/PMS system, (c) PMS dosage, (d) Fe to biochar impregnation ratio, (e) initial TCE concentration, and (f) temperature on TCE removal by Fe-CB600/PMS (pH = 8.2) (TCE concentration: 0.1 mM, catalyst dosage: 1 g L−1, PMS concentration: 5 mM, pH = 8.2, T =25 °C).

be achieved with Fe:C of 1:1 in 0.4 h (Fig. 5d). However, further increasing the Fe loading to 2:1 showed an obvious inhibition of the degradation rate, suggesting that the Fe content presented a significant impact on effective PMS activation. It is inferred that higher and lower Fe dosages were not the optimal dosage between Fe2+ and PMS, and it will hinder the activity of iron (Oh et al., 2016). When the TCE concentration was increased to 0.3 and 0.6 mM, 93.6 and 76.6% removal of TCE was obtained in 20 min (Fig. 5e), it was likely limited by the insufficient amounts of PMS or Fe/C catalyst. The TCE removal rate reached 9.80, 14.9, and 12.12 h−1 at 15, 25, and 35 °C (Fig. 5f), respectively, indicating the pathway of thermal activation of PMS (Hu et al., 2018a; Lee et al., 2012). Yet, further investigations are needed to elucidate the interplay and electron transfer between the aromatic structure of biochar matrix and the doped transition metal as a sustainable carbon-supported catalyst.

observation illustrates the important synergy of good dispersion of nZVI particles on the porous biochar support, which increased the number of active sites on the composites for the TCE removal. The removal of TCE by biochar in the absence and presence of PMS was shown in Fig. S4a. In the system of biochar alone, the removal of TCE in 20 min by SB300, CB300, SB600, and CB600 were 10.5%, 13.5%, 39.0%, and 44.0%, respectively. In comparison, the addition of PMS enhanced the TCE removal by SB300 (from 10.5%–20.0%), CB300 (from 13.5%–24.5%), SB600 (from 39% to 58.5%), and CB600 (from 44.0%–70.0%), respectively. The adsorption and oxidation processes were at work together for achieving TCE removal. Due to the best performance of Fe-CB600, the following experiments were conducted in Fe-CB600/PMS system for further investigation. The influence of experimental conditions (i.e., PMS dosages, Fe-to-biochar impregnation ratios, initial TCE concentrations, and reaction temperature) was investigated (Fig. 5). The optimal TCE removal (kobs =22.0 h−1) was achieved with 5 mM PMS (Rastogi et al., 2009). Further increasing the PMS dosage showed insignificant enhancement of the degradation rate (Fig. 5c). Only 83.3% TCE removal was reached at Fe to CB600 mass ratio (Fe:C) of 0.5:1 in 2 h, while 100% removal could

3.3. Contribution of different ROSs to PMS activation and TCE degradation The degradation of TCE can be determined by monitoring the concentration of Cl− in solution, which has been adopted and verified 6

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decomposition of PMS, and ultimately generated oxygen-containing groups on the surface of the Fe-CB600 (Zhou et al., 2015). The XPS spectra of the used Fe-CB600 on C 1s region was exhibited in Fig. 7. Compared with fresh Fe-CB600, the peak ratio of C]O functional group decreased from 13.7 to 11.2% with the concentration of C − O increased from 12.9 to 25.2% after reaction, suggesting that some of the surface C]O groups were reduced to C − O, which could be attributed to the interaction with PMS producing O2%− and 1O2 (Huang et al., 2018; Wang et al., 2016). The O 1s spectra was composed of Fe−O (530.3 eV), C]O (531.9 eV), and COe (533.5 eV) peaks. The increased propotion of Fe−O (from 18.9 to 22.5%) was attributed to the oxidization of the surface Fe. After reaction, the concentration of C]O functional group dropped from 55.6 to 50.6% and the concentration of C]O increased from 25.5 to 26.9%. XPS spectra on Fe 2p showed that after the reaction, the surface Fe0 was oxidized, and a sharp increase occurred to Fe3+ from 34.9 to 54.2%, which was consistent with the above O 1s result.

Fig. 6. Kinetics for TCE removal in presence of different ROSs scavengers by FeCB600/PMS (TCE concentration: 0.1 mM, catalyst dosage: 1 g L−1, PMS concentration: 5 mM, [EtOH] = [TBA] = 1 M, [L-Histidine] = [PBQ] =50 mM, pH = 8.2, T =25 °C).

by previous researches (Li et al., 2007; Tsai et al., 2008; Pham et al., 2009). In this study, the time-dependent increased concentration of Cl− was approximate to three-time of the decreased concentration of TCE, which inferred the complete degradation of TCE (Fig. S4b). In the light of the continuous Cl− disengaging (0.317 mM) during the 20-min reaction with Fe-CB600/PMS (Fig. S4b), the 0.1 mM TCE may have been completely removed through dechlorination via the PMS-induced ROSs and Fe0-mediated chemical reduction. However, it was difficult to differentiate the contribution of ZVI declorination and oxidation by ROSs, because TCE was dissolved in organic solvent (methanol (CH3OH)) and impossible to test the differences in solution TOC before and after reaction. Intermediate products such as formic acid, glyoxylic acid, dichloroacetic acid, and oxalic acid may be generated in the oxidative degradation process of TCE, which are eventually completely oxidized into CO2, H2O, and Cl− (Li et al., 2007). Scavenging experiments were therefore performed to discern the relative contribution of different forms of ROSs (i.e., %OH, SO4%−, O2%−, and 1O2) generated in the Fe-CB600/PMS system (Fig. 6). TBA is well known as an efficient %OH scavenger but not sensitive with SO4%− because of its higher reaction rate constant for %OH (3.8 − 7.6 × 108 M−1 s−1) and lower value for SO4%− (4.0 − 9.1 × 105 M−1 s−1) (Li et al., 2007; Tsai et al., 2008; Pham et al., 2009; Anipsitakis and Dionysiou, 2004). Owing to the high reaction rate constants with both SO4%− (1.6 − 7.7 × 107 M−1 S−1) and %OH (1.2–2.8 × 109 M−1 s−1), EtOH was then employed to quench both SO4%− and %OH (Anipsitakis and Dionysiou, 2004; Qi et al., 2013; Tang et al., 2015; Yao et al., 2016). Besides, PBQ and L-Histidine were used to scavenge the oxidation of TCE by O2%− and 1O2, respectively, in consideration of the higher reaction rate (k(O2%−) = 0.9 − 1.0 × 109 M−1 s−1, k(1O2) = 5.0 × 107 M−1 s−1) (Lee et al., 2016; Miskoski and García, 1993). The radical quenching tests indicate that the addition of %OH and SO4%− scavengers (i.e., TBA and EtOH) at a molar ratio to PMS of 200:1 had an insignificant impact on the TCE degradation with 0 and 14% decrease in the removal efficiency (Fig. 6), respectively, implying that Fe-CB600/PMS is an oxidative system not relying on %OH and SO4%−. However, the presence of PBQ (0.05 M) and L-histidine (0.05 M) significantly reduced the TCE removal efficiency to 35.9% (19.9 h−1) and 38.3% (18.4 h−1) in 0.4 h, respectively, indicating that O2%− and 1O2 in fact played more important role for TCE degradation. TEMP was used to detect the production of 1O2 at room temperature without light. A typical three-line signal of TEMPO adducts were observed (Fig. S6), which can be assigned to the oxidation of TEMP by 1 O2, and the intensity maintained during 20-min reaction. To gain insight into the TCE degradation mechanisms by the Fe/C composites, XPS investigation of the Fe-CB600 composite before and after reaction was conducted. The chemical composition analysis based on XPS full survey (Fig. 7) revealed higher O content (34.4 wt%) on the Fe-CB600 after reaction. Besides the oxidation of Fe, the catalyst may also activate and break O − O bond in PMS, resulting in the

3.4. Reusability of the Fe/C composites and TCE degradation in simulated groundwater As shown in Fig. S7a, the Fe-CB600 composite demonstrated good stability and reusability with a marginally decreased TCE degradation efficiency (from 100 to 80%) after three cycles. In contrast, there was an obvious decline in the TCE removal capacity in the fourth run, which still maintained over 50% TCE degradation. This could be attributed to the high graphitic degree, unique helical structure, and formation of abundant Fe0 nanoparticles. The spend materials can be recycled via magnetic force and regenerated through solvent extraction, which will be considered in the future research. In order to evaluate the field applicability of the PMS/Fe-CB600 system for ground water remediation, the time-dependent TCE removal was tested in the simulated groundwater under the same conditions (Supplementary Information). As shown the Fig. S8b, 90% TCE was degraded in 20 min by the Fe-CB600/PMS in simulated groundwater and the data was fitted into the pseudo-first order kinetics model (Fig. S7b). Comparing with the performance in the buffering solution (100% removal in 20 min), the TCE degradation efficiency was slightly inhibited, which indicated a negative impact of co-existing inorganic compounds in the groundwater. The price of PMS (48.2 US $/kg according to online quotation) is affordable. The Fe/C composites could be fabricated from waste biomass and Fe-containing liquid/solid wastes, which can sustainably recycle agricultural/industrial wastes, further lower the synthesis cost, and enhance the organic degradation. However, it is still premature to estimate the cost of full-scale applications based on laboratory studies. 3.5. Influences of co-existing anions The natural groundwater matrix typically contains anions such as chloride (Cl–), nitrate (NO3–), phosphate (PO43–), and hydrogen phosphate (HPO42–) that would affect the generation and transformation of sulfate radical (Zhou et al., 2015; Hu et al., 2018b). The Fig. S6 clearly depicted the influences of various anions on the TCE degradation. With the increase of Cl– concentration from 0 to 5 mM, the TCE degradation efficiency dropped from 100 to 95%, while further addition of Cl– to 20 mM significantly inhibited the performance to 89% TCE removal in 20 min. The inhibitory effect of Cl– has also been reported in previous studies (Zhou et al., 2015; Xu et al., 2016). The presence of NO3– showcased a similar effect with Cl– on the TCE degradation, whereas the coexisting PO43– and HPO42– (20 mM) induced a more remarkable restriction on the TCE removal, i.e., 40 and 62% degradation after 20min reaction, respectively. This more obvious effect was probably owing to a high quenching ability of ROSs and complexation with reactive Fe species by phosphate anions (Hu et al., 2018b; Xu et al., 2016). 7

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Fig. 7. XPS full survey and analysis of C 1s, O 1s, and Fe 2p binding states of Fe-CB600 before and after reaction with TCE/PMS.

4. Conclusions

basis of above results, the Fe-CB600 composite can serve as a highly effective and sustainable carbon-supported catalyst for organic degradation in groundwater remediation.

The biochar-supported nZVI synthesized in this study were highly efficiency for TCE degradation can be realized using PMS oxidation catalyzed by maize stalk and maize cob-derived biochar. The synthesized Fe/C composites were characterized by well-dispersed iron nanoparticles on the surface of biochar and hierarchical pores. We demonstrated that the Fe-CB600 composite exhibited a superior catalytic performance in PMS activation for TCE degradation and complete mineralization, out-performing freshly prepared nZVI and other fabricated Fe/C catalysts (i.e., Fe-SB300, Fe-SB600, and Fe-CB300). Porous biochar carrier enhances PMS activity and promotes the removal of TCE by reducing the aggregation of iron nanoparticles. The superoxide radical and singlet oxygen were validated as the predominant ROSs. On the

Acknowledgements The study was supported by National Nature Science Foundation of China (41472232, 41731282), Fundamental Research Funds for the Central Universities (2652017236), project from China Geological Survey (DD20160312), and Hong Kong Research Grants Council (PolyU 15223517).

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Appendix A. Supplementary data

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