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Environ Eng Res > Volume 28(4); 2023 > Article
Zhang, Wang, Guo, Chen, and Wang: Fabrication and Characterization of Bimetal Sulfide Hybrid Ag/S-ZVI for Dechlorination of Trichloroethylene

Abstract

Sulfidated and bimetal modification can improve the reactivity and selectivity of zero-valent iron in contaminant removal. However, previous studies were focused on unilateral material approaches. In this article, a new hybrid composite Ag/S-ZVI was prepared by doping elemental silver on surface of sulfidated nanoscale zerovalent iron (S-nZVI). The results of X-Ray Diffraction (XRD), energy dispersive spectrometer (EDS) and X-ray Photoelectron Spectroscopy (XPS) indicated Ag was dispersed on the surface of S-nZVI. Scanning electron microscope (SEM) images displayed Ag/S-ZVI was irregular clusters with rough surfaces, small agglomeration, and quasi-spherical in shape. The removal of TCE by Ag/S-ZVI was 90.4% within 20 min. The reaction rate constant Kobs of Ag/S-ZVI was 2.59 times or 1.85 times of nZVI or S-nZVI, respectively. Cycling experiments showed that Ag/S-ZVI had good recycling ability, and the removal rate of TCE reached 72.6% at the third cycle. The addition of Ag+ makes S-nZVI as an abundant and efficient source of reducing electrons. The Fe0 core can break C-Cl bonds by releasing electrons and the surface layer of Ag favors the transfer efficiency of electrons. Such study provides an efficient and robust ternary system for the remediation of TCE in groundwater.

Graphical Abstract

1. Introduction

Trichloroethylene (TCE) is a halocarbon commonly used as an industrial solvent and a wide concern as a typical pollutant. Due to the hydrophobicity, TCE enrich on the soil surface and the diffusion rate of TCE from the soil to the groundwater is extremely slow, posing a potential hazard to drinking water [1]. Many methods such as adsorption, reduction, advanced oxidation technology, and biodegradation have been adopted to degradate TCE [25]. Among various treatments, subsurface injection of nano-zero-valent iron (nZVI) for reductive dechlorination is a simple, efficient, cost effective, and environmentally friendly alternative [6]. Meanwhile, due to the magnetic properties of iron, iron-based nanomaterials are easily separated from treated water, which not only avoids secondary pollution of water, but also saves resources and facilitates recycling [7].
Due to the strong electronegativity of nZVI, it can provide electrons in the reductive dechlorination reaction. However, the nanoparticles tend to clump together and become more stable and less reactive [8,9]. It would limit the application of nZVI, so modification of nZVI is focused on [10]. Among the methods to improve the performance of nZVI, sulfidated zero-valent iron (S-ZVI) has been widely noticed as an emerging modification method [11]. The introduction of chelating agent leads to the change of nZVI surface charge, prevents electrostatic attraction between nZVI particles and reduces their aggregation through electrostatic stabilization effect and improves the stability of nZVI. The removal of TCE by S-nZVI was much higher than that of nZVI, and the reaction mechanism mainly occurred by reductive dechlorination [12,13]. Sulfidation can greatly increase the overall electron efficiency or selectivity of reducing contaminants [14]. The S-nZVI nanoparticles improve the electron selectivity of the material due to the generation of FeSx which hinder the agglomeration of nZVI and inhibit the reaction of Fe0 with water. At the same time, FeSx is hydrophobic and can maintain the high activity of nanoparticles. Therefore, the modification of ZVI nanoparticles with sulfide was chosen for the removal of TCE [15].
Compared to nZVI, addition of another metal to form bimetallic nanoparticles is another way to improve the reactivity of nZVI or ZVI. Both atomic hydrogen activation and electron transfer play important roles in the TCE reduction process [16,17]. The reduction potential (E0) of metal ions relative to that of Fe0 is an important factor affecting the removal effect of TCE [18]. Those metals that have more positive E0 values than Fe0, such as Pd2+, Cu2+ and Ag+, who are readily reduced to the corresponding zero-valent metals, forming iron-based bimetals. As soon as bimetals are formed, they can enhance the reaction rate by electrostatic coupling action between both metals [1921]. The dechlorination rate of iron-based bimetallic nanoparticles was several orders of magnitude higher than that of nZVI particles under conditions of oxygen deficiency and iron excess [22].
Although, bimetallic system or sulfidated zero-valent iron can effectively promote the degradation of pollutants, but it does not improve the easy agglomeration and sedimentation of nZVI particles. Reducing the agglomeration between nZVI and enhancing the colloidal stability of nZVI is another hot spot. Therefore, the reactivity of the material was further improved by trying to combine two more promising means. However, to our knowledge, till now studies on bimetallic modified S-nZVI have still rarely been reported. Since Ag has a high standard potential (0.799 V), the galvanic couple formed by Fe (E0Fe (II)/Fe = −0.44 V) and Ag has a high potential and a fast electron transfer rate [23,24]. Therefore, in order to further improve the removal efficiency of TCE, the modification of S-nZVI nanoparticles with addition of Ag+ ought to be attempted.
In this study, composite of Ag/S-ZVI was prepared. Batch experiments were determined by gas chromatography flame ionization detector (GC-FID) headspace method to explore the effects of different paraments on the degradation rate of TCE. The recyclability of the Ag/S-ZVI was verified by with nZVI, S-nZVI through cyclic experiments. In addition, the reaction process was fitted with kinetic analysis to investigate the mechanism of adsorption and dechlorination of TCE by the composite. By combining the both modifications, we hope to improve the TCE removal and recycling ability of the S-nZVI, and to investigate the mechanism of TCE removal by the Ag/S-ZVI through characterization analysis and experimental results.

2. Materials and Methods

2.1. Chemicals and Materials

All chemicals used in this study, including sodium borohydride (NaBH4), ferrous sulfate (FeSO4·7H2O), sodium sulfide (Na2S·9H2O), silver nitrate, methanol, trichloroethylene, sodium hydroxide, and sulfuric acid, were of analytic grade and purchased from Shanghai Titan Scientific Co., Ltd. In all experiments, the deionized water (pH=6.0 without adjustment) was first pretreated under N2 atmosphere to avoid oxidation during the reaction.

2.2. Synthesis of Ag/S-ZVI

S-nZVI was first prepared by liquid-phase reduction, and then Ag/S-ZVI composites were prepared by loading Ag onto it through the substitution reaction. All processes were carried out under N2 atmosphere to avoid the oxidation of Fe2+ or Fe0. Typical process was adopted as follows. 0.249 g FeSO4·7H2O was dissolved in 25 mL water. A 20 mL mixed solution containing NaBH4 and Na2S·9H2O was added dropwise into the Fe (II) solution. After 30 min, 3.9 mM AgNO3 solution was added dropwise into the S-nZVI solution and then reacted for another 30 min. The synthetic process of Ag/S-ZVI was shown in Fig. S1.
The solid residue was washed with water and ethanol and dried in vacuum oven at 60°C to give resulting composite for further use.

2.3. Characterizations and Performance Testing

The crystal structure of samples was analyzed using X-ray diffractometer (Bruker D8-Advance, Germany). Morphologies of samples were examined by scanning electron microscopy (SEM, Hitachi S-4800, Japan) with energy dispersive spectrometer (EDS) for analyzing the surface elemental content. The Brunauer-Emmett-Teller (BET) surface area of the material was determined by N2 adsorption desorption isotherm (MicritracBel belsorp-max). X-ray photoelectron spectrometer (XPS, Thermo Scientific K-Alpha) was used to characterize elements states and chemical composition.

2.4. TCE Removal: Batch Experiments

Catalysts and prepared TCE solution were added into 100 mL serum bottles, which were sealed by rubber plugs. The 1 mL of sample was withdrawn at fixed time points using a micro injection. The concentration of the extracted TCE was measured by gas chromatography (GC-FID), and two parallel experiments were set up for each determination [25,26]. The peak time of methanol and TCE was 3.6 and 4.6 min, respectively. Peak areas of different concentrations of TCE were measured by GC-FID, and the standard curves were plotted according to the relationship between peak areas and concentrations. The standard curve of the experiment was obtained as Fig. S2.
The TCE removal rate have been calculated as Eq. (1).
(1)
R%=Co-CeCo×100%
where R is the removal rate of TCE, C0 (mg /L) is the initial TCE concentration, and Ce (mg /L) is the final TCE concentration in the treated solution.
The one-order kinetic fitting was investigated in order to study the reaction rate for TCE removal. The fitted model is expressed as Eq. (2).
(2)
-ln(C/C0)=Kobst
where Kobs is the reaction rate constant, min−1; C0 (mg /L) is the initial TCE concentration, and C (mg/L) is the current TCE concentration in the treated solution.

3. Results and Discussion

3.1. Characterization of Ag/S-ZVI

The composition and crystallographic structures of nZVI, S-nZVI, Ag/S-ZVI were determined by XRD, as shown in Fig. 1. For Ag/S-ZVI, only weak peaks at 38.11°, 44.67°, 64.42°, and 77.47° were observed to match well with {111}, {200}, {220}and {311} diffraction planes of Ag (PDF-#87-0719). Besides, both peaks at 44.67° and 64.42° seems to be coincident with the {110} and {220} diffraction planes of Fe (PDF-#87-0722). Diffraction peaks observed in the XRD pattern of Ag/S-ZVI are demonstrating the coexistence of Ag [27] and Fe. The diffraction peaks of sulfur-containing compounds such as FeS or FeS2 are not detected, which may be due to their low concentration or low crystallinity. The XRD patterns of the nZVI and S-nZVI show typical diffraction peaks, thus implying the successful preparation of nZVI and S-nZVI. The effect of the Fe/Ag molar ratio on the material structure could be seen in Fig. 1. The crystallinity of the material is relatively decreased with the more addition of Ag.
The morphology of Ag/S-ZVI was carried out by SEM, as shown in Fig. 2. The SEM micrograph clearly shows irregular clusters with rough surfaces, small agglomeration, and quasi-spherical in shape. The formed silver nanoparticles are clustered and unevenly overlapped on surface of S-nZVI.
To gain more facts about the composition and elements in Ag/S-ZVI, energy dispersive spectrometer (EDS) analysis was carried out. The result of EDS was shown in Fig. S3. The peaks of Ag and Fe nanoparticles appear at 3 keV (Ag), 6.6 keV (Fe) and 0.9 keV (Fe), further demonstrating the presence of Ag and Fe. The intensity of the Fe signal in the EDS spectrum was higher than that of the Ag signal, resulting in a high density of iron in the inside region. The presence of O peak in EDS spectra might be due to the air in the SEM grid formation. EDS elemental mapping images were also presented in Fig. S4. Based on the O–K, S–K, Fe-L and Ag-L signals and their distribution area, Ag and S are dispersed on top of nZVI [28,29].
N2 adsorption-desorption isotherm and Barrett-Joyner-Halenda (BJH) curve of Ag/S-ZVI were shown in Fig. S5, respectively. The specific surface area, average pore size, and pore volume of prepared materials were collected in Table S1. The specific surface area of Ag/S-ZVI was 74.9 m2/g that was higher than those of nZVI (49.0 m2/g) and S-nZVI (64.6 m2/g). The more specific surface area of Ag/S-ZVI can provide, the more active sites and contact area were for the degradation of pollutants, thus improving the degradation efficiency.
The elemental composition and valence information derived from the X-ray Photoelectron Spectroscopy (XPS) characterization analysis of the Ag/S-ZVI was shown in Fig. S6. The peaks of the four elements O, Fe, S and Ag are clearly seen from Fig. S6a, where the peak of C 1s was caused from the introduction of exogenous C for a standard. The peak at 719.6 eV corresponding to Fe0 is evident from the Fe 2p spectrum (Fig. S6b) [30]. FeO, Fe2O3 peaks were located at 709.8 eV, 723.9 eV respectively. Peaks of 711.5 eV and 724.5 eV were correspond to FeOOH [30,31]. The appearance of FeOOH was due to the samples used for characterization that have been dried or oxidized during the detection process to form an oxide shell, while the oxide shell protects the internal Ag/S-ZVI from further oxidation to some extent. The Fe0, FeO, Fe2O3, and FeOOH relative contents of the surface are 14.25%, 35.05%, 27.83%, and 22.87%, respectively.
In Fig. S6c, the most prominent peak S2− was observed in the S 2p spectrum with a binding energy of 161.2 eV [32]. In addition, peaks of SO32− were also detected in the sample as S2− oxidation products [33]. The relative contents of S2−, SO32− and SO42− in Ag/S-ZVI are 38.5%, 6.62%, and 37%, respectively.
By XPS, Ag was in three forms (Ag0, Ag2O, Ag2S) in Fig. S6d. The peaks of Ag were at 371 and 374.2 eV and the peak of Ag2O was at 368.2 eV (which also contains Ag singlet according to the binding energy spectrum). The peak of Ag2S was at 366.8 eV. The Ag, Ag2O, Ag2S relative contents were 36.74%, 59.63%, and 3.63%. In synthesis process, the addition of Ag+, Fe (II) undergoes a substitution reaction with Ag+ to produce Ag0. Reduction and deposition of silver on the surface of S-nZVI occurred through Eq. (3) and Eq. (4) [34]. In the reaction solution, due to the instability of metallic silver, an outer nonuniform layer of Ag is formed on the surface of the S-nZVI [35,36]. The presence of Ag2O is attributed to that Ag was oxidized during the subsequent recovery and treatment of drying. Ag2S is generated by the reaction of FeSx on the surface of S-nZVI during the addition of AgNO3. At higher concentrations of sulfur in solution, a small amount of monomeric Ag is converted directly to Ag2S through a solid-liquid reaction [35,36].
(3)
Fe2++BH4-+H2OFe0+S2-S-nZVI
(4)
2Ag++Fe=Fe2++2Ag

3.2. TCE Removal by Ag/S-ZVI

3.2.1. Influence of different materials

Compared to nZVI, S-nZVI and Ag-nZVI, the removal rate of Ag/S-ZVI for the removal of TCE was shown in Fig. 3a. In the first 20 min, the removal percentage of TCE by nZVI, S-nZVI, Ag-nZVI and Ag/S-ZVI was 56.4%, 73.5%, 70.6% and 90.4%, respectively. Obviously, Ag/S-ZVI can degrade TCE at a higher rate than other three.
The Kobs of different materials were further investigated by fitting the experimental results with one-order kinetic. The fitting results were shown in Fig. 3b, and fitting parameters were shown in Table S2. The reaction rate constant Kobs of Ag/S-ZVI was significantly higher than that of nZVI, S-nZVI and Ag-nZVI, which were 2.59, 1.85, 1.97 times higher. For nZVI, it is very easy to agglomerate and be oxidized, which make less contact with pollutants and restricts its continued reaction. Due to the poor of electron selectivity of nZVI, it is easy to react with dissolved oxygen in water, which reduces removal of target pollutants [37].
After sulfidation, the resistance to oxidation is improved. The generated sulfide ferrite hinders the agglomeration of nZVI, improving the electron selectivity. Since Ag has a high standard potential (0.799 V), the galvanic couple formed by Fe (E0Fe (II)/Fe = −0.44 V) and Ag has a high potential and a fast electron transfer rate. The surface area normalized rate constants (ksa) of the three materials were shown in Table S2. With the addition of Ag+, the surface area normalization rate constant of Ag/S-ZVI (1.5×10−3, g·m−2·min−1) was significantly higher than those of nZVI (8.6×10−4, g·m−2·min−1), S-nZVI (9.2×10−4, g·m−2·min−1) and Ag-nZVI (9.0×10−4, g·m−2·min−1). The surface layer of FeSx is negatively charged and can attract or adsorb a high density of Ag+. At the same time, Ag-Fe0 structure endows the composite with an abundant and more efficient source of reducing electrons [38]. The Fe0 core can break C-Cl bonds by increasing the transfer efficiency of electrons through the surface layer of Ag. In addition, the enhanced adsorption capacity of Ag/S-ZVI increases the contact of TCE with Ag/S-ZVI.

3.2.2. Influence of S/Fe molar ratio

The highest TCE removal rate of 90.4% was achieved when the S/Fe molar ratio is 0.5 from Fig. 3c. The electronic selectivity is enhanced to improve the reduction ability of Ag/S-ZVI for TCE. FeSx manifestation affects the hydrophobicity of S-nZVI [39]. From Fig. 3c, it can be seen that the S/Fe molar ratio was 0.2 and the removal rate of TCE was 85.9%. When S/Fe molar ratio was low, the FeSx on the surface was almost FeS2 with higher hydrophobicity [39], which was not conducive to the introduction of Ag+. When the S/Fe molar ratio reached 0.5, the removal rate of TCE was 90.4%, sulfur was present in the form of FeS. With the content of sulfur increase, the reactive sites on the surface of Ag/S-ZVI are blocked by the excess of FeS and Ag, which was not beneficial to reactions of material in contact with TCE. When S/Fe molar ratio reached 0.8, the removal rate of TCE was 79.6%.
The sulfide ferrite surface layer of Ag/S-ZVI acts as an effective electron conductor, which helps electrons transfer from Fe0 core to the surface. The FeS shell of S-nZVI is a chemically heterogeneous and defective structure, which offers efficient electron transfer and high reactivity via tunneling effects and/or defect channels [40,41]. In addition, Ag/S-ZVI has increased surface roughness and surface area, improving the reactivity of system [25].

3.2.3. Influence of Ag+ addition

The addition of Ag+ also acts as a key factor on the removal of TCE. The highest removal rate of TCE was 90.4% when 3.9 mM Ag+ was involved in Fig. 3d. And after sulfidation, the sulfur atoms preempt some reaction sites. With the addition of a small amount of Ag, the silver atoms occupy the remaining part of the reactive sites, which facilitates the synergistic effect of the composite on TCE removal. However, the degradation rate of TCE gradually decreased from 90.4% to 76.9% when the addition amount of Ag+ was more than 3.9 mM. With the addition of excess Ag, the contact between the material and TCE is reduced, which in turn decreases the TCE removal rate.

3.2.4. Influence of initial pH

The influence of initial solution pH on the removal of TCE by Ag/S-ZVI was shown in Fig. 3e. The solution pH was adjusted from 2.0 to 10.0 by adding H2SO4 or NaOH. The optimal solution pH for TCE removal by Ag/S-ZVI is 6.0 and the removal rate was 90.4%. When the solution is at pH 6.0, the surface layer corrosion is favored, and electrons are released to reduce TCE. The surface passivated layer of Ag/S-ZVI is highly active after corrosion and can reduce more TCE. At pH 6.0, the shell will gradually corrode away to maintain the reduction rate of S-nZVI, so the reaction efficiency will be increased. At pH 2.0 or pH 4.0, the shell layer is gradually corroded and Fe0 is consumed to form a large amount of Fe2+, which does not have the ability to reduce TCE and the reaction efficiency decreased instead. However, the TCE removal rate decreases with the increase of pH value from pH 6.0 to pH 10.0. It was shown that less hydrogen is produced, and more passivation layers formed on the surface under alkaline conditions [42]. The point of zero charge (pHPZC) of Ag/S-ZVI was determined from pH 2.0 to 10.0. The result was shown in Fig. S7. The pHPZC of the Ag/S-ZVI was at about 6.6, indicating that when pH < pHPZC, the Ag/S-ZVI has positive surface charge and can attract anions. Therefore, the acidic conditions promote the release of electrons from Fe0 corrosion and eventually TCE is reduced by electrons.

3.2.5. Influence of initial TCE concentration

The result of initial concentrations of TCE (10–90 mg/L) on removal rate by Ag/S-ZVI was studied and shown in Fig. 3f. It can be seen that the lower the initial concentration of TCE, the better the degradation of TCE by Ag/S-ZVI.
A pseudo primary and pseudo secondary kinetic model for the degradation of TCE in aqueous solution by Ag/S-ZVI was developed. The pseudo-primary and pseudo-secondary kinetic models are Eq. (5) and Eq. (6).
(5)
ln(qe-qt)=ln qt-K1t
(6)
tqt=1K2qe2+tqe
where K1 is the pseudo primary adsorption rate constant, min−1; K2 is the pseudo secondary adsorption rate constant, g·mg−1·min−1; qe is the equilibrium adsorption amount, mg/g; qt is the adsorption amount at time t, mg/g.
The fitting results were shown in Figs. 3g–3h and fitting parameters were collected in Table S3. Apparently, the fitted results are consistent with the pseudo-second-order kinetic model (R2 > 0.99). The removal of TCE by Ag/S-ZVI is mediated by both adsorption modes from the above result, but mainly by chemisorption [43]. This process may involve two steps: (1) sorption of TCE to samples surface, (2) reductive degradation or reduction of TCE on surface. The adsorption capacity is limited at a fixed adsorbent concentration. As the initial concentration of TCE gradually increases, the competitive adsorption among TCE will gradually increase.

3.2.6. Cycling experiments of Ag/S-ZVI

Reusability of composite materials is a key criterion in evaluating whether the material has practical applications. By comparing cycling experiments of nZVI and S-nZVI, the cycling performance of Ag/S-ZVI is significantly better than the other two materials from Fig. S8. At the completion of the first set of cyclic experiments, the used Ag/S-ZVI was separated by centrifugation and washed with deoxidation water and ethanol, and then proceeded to the next cycle. Combining the three materials, the number of TCE degradation cycles was tested three times, and the removal rates during different cycles were shown in Fig. S8. Ag/S-ZVI still reached 72.6% for TCE removal rate at the third run, which is much higher than S-nZVI (33.6%) and nZVI (18.8%). Inductively coupled plasma atomic emission spectroscopy (ICP-AES) was used to analysis the concentration of silver ions after the reaction. If the dosage of silver in the Ag/S-ZVI (Fe/Ag molar ratio=20:1) is dissolved in the solution, the concentration of silver ions would be 48 mg/L. The final silver ion concentration of the reaction was 0.03 mg/L from ICP-AES, and the silver ion leaching rate is 0.0625%. The results showed that the amount of silver ions leached was quite low, which was conducive to the recycling of the material. Ag on surface of S-nZVI prevents further oxidation of S-nZVI and ensure sustainable reuse of the material. And sulfur ferrites also help to promote electron transfer between the iron ions and contaminants as well as to promote electron transfer from the Fe0 core to its surface.

3.2.7. Mechanisms of TCE removal

For Ag/S-ZVI, reduction should be the main mechanism. It can be raised that the introduction of Ag plays a facilitating role in the reduction of TCE by S-nZVI. Consequently, considerable reaction activities of nZVI are obtained.
To better understand the TCE removal by the Ag/S-ZVI, the variations in phase and surface valence state of composite were analyzed utilizing SEM and XPS.
The SEM of Ag/S-ZVI materials before and after the reaction was shown in Fig. 4. It can be seen from SEM images that spherical structures of Ag/S-ZVI after the reaction are not obvious and starts to rough up to form blocks. During this process, the shell of sulfide was corroded during the removal of TCE and also inhibits the continuation of the reaction. The Fourier Transform infrared spectroscopy (FTIR) results of the material before and after reaction were shown in Fig. S9. From the Fig. S9, we can see that the C=C stretching vibration peak was shifted from 1636 cm−1 to 1651 cm−1. The result explained indirectly that the removal of TCE was a reduction process rather than an adsorption on the material surface. In order to further investigate the valence differences of the elemental contents, further XPS analysis was performed.
In Fig. 5, the Fe 2p peak shows that the Fe0 content of Ag/S-ZVI remained as 8.9% after the reaction compared to the Fe content before the reaction, which indicates that Ag/S-ZVI reduces TCE by electron release from Fe0 corrosion, and the remaining Fe0 retained by Ag/S-ZVI can be left to continue the reaction cycle. The increase in FeO content may be due to the reaction that S-nZVI reacts with hydrogen ions (H+) to form iron oxides or hydroxides from the reaction of some of the iron ions generated and OH in solution.
S 2p fractionation peak was shown in Fig. 6. The amount of S2− detected in the Ag/S-ZVI after the reaction decreases from 38.50% to 25.72%, while the amount of SO42− detected in the surface increased from 37.0% to 42.77%. This indicates that Ag/S-ZVI consumes a certain amount of Fe0/FeS during the degradation of TCE, so that part of S2− also participates in the reaction to produce SO42−. This proves that the sulfur ferrite shell is corroded during the reaction by participating in the reaction process, thus confirming the phenomenon reflected by the SEM images.
The characteristic spectral lines of Ag 3d before and after the reaction are almost the same, indicating that silver remains essentially unchanged as a catalyst. However, the peak area of the Ag 3d region decreases from 1641.28eV to 1335.81eV before and after the reaction, indicating that the concentration of Ag decreases during the reaction. By contacting the material in the solution, the protons on the Ag surface are transformed into adsorbed atomic hydrogen by the Volmer reaction (H+ + e + Ag → Ag-H*) and then dissociated into more active atomic hydrogen by electron transfer [24]. Then atomic hydrogen acts on the C-Cl bond to reduce it to C-H bond and chloride ion by a surface contact process [34]. With a thin discontinuous layer of Ag, the Fe0 core can also provide electron to break the C-Cl bond [44].
The reaction with Ag/S-ZVI is a surface-controlled electrochemical reaction in which Fe0 provides electrons and is oxidized to Fe2+ or Fe3+. While trichloroethylene gains electrons, the C-Cl bonds were broken, and the chlorine atoms are shed to form chloride ions. Therefore, the presence of chloride ions in the solution after the reaction is the key factor to determine whether the dichlorination of TCE occurs. The ion chromatograph (Dionex ICS-2100) was used to detect the concentration of chloride ions after the reaction, and the results were shown in Fig. S10. The theoretical Cl concentration would be 39.9 mg/L if three chlorine atoms have been dechlorinated. The results showed that the Cl concentration in the solution reached 26.8 mg/L after the reaction was completed.
The reaction mechanism of Ag/S-ZVI for TCE removal was described in Fig. 8. The Fe0 core is an electron source to provide electron and the sulfide ferrite surface layer helps electrons transfer to the surface as an effective electron conductor. Ag acts as a catalyst to accelerate electron transfer to generate more atomic hydrogen, which is used to reduce the C-Cl bond. Ag-Fe0 structure endows the composite with an abundant and more efficient source of reducing electrons. Theoretically, the removal process of TCE was as Eq. (7) and Eq. (8).
(7)
Fe0Fe2++2e-
(8)
C2HCl3+ne-+(n-3)H+Products+3Cl-
By combining the two modifications, we improve the TCE removal and recycling ability of the S-nZVI. Because of these advantages, we hope Ag/S-ZVI become a potent alternative to traditional treatment methods. However, there are still limitations and shortcomings. Due to the limitation of experimental conditions and equipment, all the simulated wastewater was used to replace the polluted wastewater in this study, and the complexity of pollutants and the generation of combined reactions in the actual wastewater were not considered. For the study of Ag/S-ZVI, it is only a laboratory scale, and field scale experiments will be further investigated.

4. Conclusions

In this study, Ag/S-ZVI was synthesized, characterized, and investigated to remove TCE from aqueous solution. S-nZVI was synthesis by a one-step synthesis method and then the composite was produced by addition of Ag+. The results of SEM and EDS showed that Ag was successfully dispersed on the surface of S-nZVI. Ag/S-ZVI clearly shows irregular clusters with rough surfaces, small agglomeration, and quasi-spherical in shape by SEM. Batch experiments indicated that Ag/S–nZVI was more efficient for TCE removal than S-nZVI or nZVI. The removal of TCE by Ag/S-ZVI reached 90.4% within 20 min (S/Fe molar ratio =0.5, Ag+ addition =3.9mM, pH=6). The reaction rate constant Kobs of Ag/S-ZVI was 2.59 times or 1.85 times higher than that of nZVI or S-nZVI, respectively. The calculations of Kobs further demonstrated that Ag/S-ZVI was ideal to remove TCE. Good recyclability of Ag/S-ZVI was evaluated by cycling experiments. The addition of Ag accelerates electron transfer while generating more atomic hydrogen, which is used to reduce the C-Cl bond. Ag/S-ZVI combines excellent reactivity and recyclability, making it a potential ideal material for groundwater contamination remediation applications.

Supplementary Information

Acknowledgements

This work was founded by the National Natural Science Foundation of China (No. 52070127).

Notes

Conflict of Interest Statement

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Author contributions

Z.S.B. (M.Sc. student) conducted the experiments and wrote the original draft. W.T.X. (M.Sc. student) assisted in conducting experiments. G.X. (M.Sc. student) assisted in conducting experiments. C.S.W. (Associate Professor) did Supervision, Conceptualization, Writing & Editing. W.L.J. (Professor) reviewed the whole work and provided suggestions for the improvement of the paper.

References

1. Mohammad OI, Kibbey TCG. Dissolution-Induced Contact Angle Modification in Dense Nonaqueous Phase Liquid/Water Systems. Environ. Sci. Technol. 2005;39(6)1698–1706. https://doi.org/10.1021/es048810o
crossref pmid

2. Khan MS, Shah JA, Riaz N, et al. Synthesis and Characterization of Fe-TiO2 Nanomaterial: Performance Evaluation for RB5 Decolorization and In Vitro Antibacterial Studies. Nanomaterials. 2021;11(2)436 https://doi.org/10.3390/nano11020436
crossref pmid pmc

3. Sarwar A, Wang J, Khan MS, et al. Iron Oxide (Fe3O4)-Supported SiO2 Magnetic Nanocomposites for Efficient Adsorption of Fluoride from Drinking Water: Synthesis, Characterization, and Adsorption Isotherm Analysis. Water. 2021;13(11)1514 https://doi.org/10.3390/w13111514
crossref

4. Khan MS, Riaz N, Shaikh AJ, et al. Graphene quantum dot and iron co-doped TiO2 photocatalysts: Synthesis, performance evaluation and phytotoxicity studies. Ecotoxicol. Environ. Saf. 2021;226:112855 https://doi.org/10.1016/j.ecoenv.2021.112855
crossref pmid

5. Khan MS, García MF, Javed M, et al. Synthesis, Characterization, and Photocatalytic, Bactericidal, and Molecular Docking Analysis of Cu–Fe/TiO2 Photocatalysts: Influence of Metallic Impurities and Calcination Temperature on Charge Recombination. ACS Omega. 2021;6(40)26108–26118. https://doi.org/10.1021/acsomega.1c03102
crossref pmid pmc

6. Fu F, Dionysiou DD, Liu H. The use of zero-valent iron for groundwater remediation and wastewater treatment: A review. J. Hazard. Mater. 2014;267:194–205. https://doi.org/10.1016/j.jhazmat.2013.12.062
crossref pmid

7. Al-Anazi A. Iron-based magnetic nanomaterials in environmental and energy applications: a short review. Curr. Opin. Chem. Eng. 2022;36:100794 https://doi.org/10.1016/j.coche.2022.100794
crossref

8. Xu J, Sheng T, Hu Y, Baig SA, Lv X, Xu X. Adsorption-dechlorination of 2,4-dichlorophenol using two specified MWCNTs-stabilized Pd/Fe nanocomposites. Chem. Eng. J. 2013;219(1)162–173. https://doi.org/10.1016/j.cej.2013.01.010
crossref

9. Li J, Zhang X, Sun Y, et al. Advances in Sulfidation of Zerovalent Iron for Water Decontamination. Environ. Sci. Technol. 2017;51(23)13533–13544. https://doi.org/10.1021/acs.est.7b02695
crossref pmid

10. Schöftner P, Waldner G, Lottermoser W, Stöger-Pollach M, Freitag P, Reichenauer TG. Electron efficiency of nZVI does not change with variation of environmental parameters. Sci. Total. Environ. 2015;535:69–78. https://doi.org/10.1016/j.scitotenv.2015.05.033
crossref pmid

11. Zhou L, Li Z, Yi Y, Tsang EP, Fang Z. Increasing the electron selectivity of nanoscale zero-valent iron in environmental remediation: a review. J. Hazard. Mater. 2022;421:126709 https://doi.org/10.1016/j.jhazmat.2021.126709
crossref pmid

12. Mangayayam M, Dideriksen K, Ceccato M, Tobler DJ. The Structure of Sulfidized Zero-Valent Iron by One-Pot Synthesis: Impact on Contaminant Selectivity and Long-Term Performance. Environ. Sci. Technol. 2019;53(8)4389–4396. https://doi.org/10.1021/acs.est.8b06480
crossref pmid

13. Fan D, Johnson G, Tratnyek PG, Johnson RL. Sulfidation of nano zerovalent iron (nZVI) for improved selectivity during in-situ chemical reduction (ISCR). Environ. Sci. Technol. 2016;50(17)9558–9565. https://doi.org/10.1021/acs.est.6b02170
crossref pmid

14. Gu Y, Wang B, He F, Bradley MJ, Tratnyek PG. Mechanochemically sulfidated microscale zero valent iron: pathways, kinetics, mechanism, and efficiency of trichloroethylene dechlorination. Environ. Sci. Technol. 2017;51(21)12653–12662. https://doi.org/10.1021/acs.est.7b03604
crossref pmid

15. Kim EJ, Kim JH, Azad AM, Chang YS. Facile Synthesis and Characterization of Fe/FeS Nanoparticles for Environmental Applications. ACS Appl. Mater. Interfaces. 2011;3(5)1457–1462. https://doi.org/10.1021/am200016v
crossref pmid

16. Kim EJ, Kim JH, Chang YS, Turcio-Ortega D, Tratnyek PG. Effects of Metal Ions on the Reactivity and Corrosion Electrochemistry of Fe/FeS Nanoparticles. Environ. Sci. Technol. 2014;48(7)4002–4011. https://doi.org/10.1021/es405622d
crossref pmid

17. Su Y, Lowry GV, Jassby D, Zhang Y. Sulfide-Modified NZVI (S-NZVI): Synthesis, Characterization, and Reactivity. Phenrat T, Lowry GV, editorsNanoscale Zerovalent Iron Particles for Environmental Restoration. 1st edSpringer; Cham: 2019. p. 359–386. https://doi:10.1007/978-3-319-95340-3_9
crossref

18. Li X, Zhang W. Sequestration of Metal Cations with Zerovalent Iron Nanoparticles-A Study with High Resolution X-ray Photoelectron Spectroscopy (HR-XPS). J. Phys. Chem. C. 2007;111(19)6939–6946. https://doi.org/10.1021/jp0702189
crossref

19. Tee YH, Bachas L, Bhattacharyya D. Degradation of Trichloroethylene by Iron-Based Bimetallic Nanoparticles. J. Phys. Chem. C. 2009;113(22)9454–9464. https://doi.org/10.1021/jp809098z
crossref pmid pmc

20. Kim JH, Tratnyek PG, Chang YS. Rapid Dechlorination of Polychlorinated Dibenzo-p-dioxins by Bimetallic and Nanosized Zerovalent Iron. Environ. Sci. Technol. 2008;42(11)4106–4112. https://doi.org/10.1021/es702560k
crossref pmid

21. Xie Y, Cwiertny DM. Chlorinated solvent transformation by palladized zerovalent iron: mechanistic insights from reductant loading studies and solvent kinetic isotope effects. Environ. Sci. Technol. 2013;47(14)7940–7948. https://doi.org/10.1021/es401481a
crossref pmid

22. He F, Li Z, Shi S, et al. Dechlorination of excess trichloroethene by bimetallic and sulfidated nanoscale zero-valent iron. Environ. Sci. Technol. 2018;52(15)8627–8637. https://doi.org/10.1021/acs.est.8b01735
crossref pmid

23. O'Carroll D, Sleep B, Krol M, Boparai H, Kocur C. Nanoscale zero valent iron and bimetallic particles for contaminated site remediation. Adv. Water. Resour. 2013;51:104–122. https://doi.org/10.1016/j.advwatres.2012.02.005
crossref

24. Luo S, Yang S, Wang X, Sun C. Reductive degradation of tetra-bromobisphenol A over iron–silver bimetallic nanoparticles under ultrasound radiation. Chemosphere. 2010;79(6)672–678. https://doi.org/10.1016/j.chemosphere.2010.02.011
crossref pmid

25. Kim EJ, Murugesan K, Kim JH, Tratnyek PG, Chang YS. Remediation of Trichloroethylene by FeS-Coated Iron Nanoparticles in Simulated and Real Groundwater: Effects of Water Chemistry. Ind. Eng. Chem. Res. 2013;52(27)9343–9350. https://doi.org/10.1021/ie400165a
crossref

26. Coopman V, Cordonnier J, De Letter E, Piette M. Tissue distribution of trichloroethylene in a case of accidental acute intoxication by inhalation. Forensic Sci. Int. 2003;134(2–3)115–119. https://doi.org/10.1016/S0379-0738(03)00131-2
crossref pmid

27. Alzahrani SA, Malik MA, Al-Thabaiti SA, Khan Z. Seedless synthesis and efficient recyclable catalytic activity of Ag@Fe nanocomposites towards methyl orange. Appl. Nanosci. 2018;8(3)255–271. https://doi.org/10.1007/s13204-018-0699-7
crossref

28. Malik MA, Alshehri AA, Patel R. Facile one-pot green synthesis of Ag-Fe bimetallic nanoparticles and their catalytic capability for 4-Nitrophenol reduction. J. Mater. Res. Technol. 2021;12(4)455–470. https://doi.org/10.1016/j.jmrt.2021.02.063
crossref

29. Ling L, Tang C, Zhang W. Visualization of Silver Nanoparticle Formation on Nanoscale Zero-Valent Iron. Environ. Sci. Technol. Lett. 2018;5(8)520–525. https://doi.org/10.1021/acs.estlett.8b00259
crossref

30. Gong Y, Gai L, Tang J, Fu J, Wang Q, Zeng EY. Reduction of Cr (VI) in simulated groundwater by FeS-coated iron magnetic nanoparticles. Sci. Total. Environ. 2017;595:743–751. https://doi.org/10.1016/j.scitotenv.2017.03.282
crossref pmid

31. Liang L, Li X, Guo Y, Lin Z, Liu B, Su X. The removal of heavy metal cations by sulfidated nanoscale zero-valent iron (S-nZVI): The reaction mechanisms and the role of sulfur. J. Hazard. Mater. 2021;404:124057 https://doi.org/10.1016/j.jhazmat.2020.124057
crossref pmid

32. Yang L, Gao J, Liu Y, et al. Removal of Methyl Orange from Water Using Sulfur-Modified nZVI Supported on Biochar Composite. Water. Air. Soil Pollut. 2018;229(11)355 https://doi.org/10.1007/s11270-018-3992-x
crossref

33. Qu J, Liu Y, Cheng L, et al. Green synthesis of hydrophilic activated carbon supported sulfide nZVI for enhanced Pb (II) scavenging from water: Characterization, kinetics, isotherms and mechanisms. J. Hazard. Mater. 2021;403:123607 https://doi.org/10.1016/j.jhazmat.2020.123607
crossref pmid

34. Wang Z, Huang W, Peng P, Fennell DE. Rapid dechlorination of 1,2,3,4-TCDD by Ag/Fe bimetallic particles. Chem. Eng. J. 2015;273:465–471.
crossref

35. Levard C, Hotze EM, Colman BP, et al. Sulfidation of Silver Nanoparticles: Natural Antidote to Their Toxicity. Environ. Sci. Technol. 2013;47(23)13440–13448. https://doi.org/10.1021/es403527n
crossref pmid pmc

36. Levard C, Hotze EM, Lowry GV, Brown GE. Environmental Transformations of Silver Nanoparticles: Impact on Stability and Toxicity. Environ. Sci. Technol. 2012;46(13)6900–6914. https://doi.org/10.1021/es2037405
crossref pmid

37. Du J, Bao J, Lu C, Werner D. Reductive sequestration of chromate by hierarchical FeS@Fe0 particles. Water Res. 2016;102:73–81. https://doi.org/10.1016/j.watres.2016.06.009
crossref pmid

38. Ling L, Zhang W. Visualizing Arsenate Reactions and Encapsulation in a Single Zero-Valent Iron Nanoparticle. Environ. Sci. Technol. 2017;51(4)2288–2294. https://doi.org/10.1021/acs.est.6b04315
crossref pmid

39. Xu J, Li H, Lowry GV. Sulfidized Nanoscale Zero-Valent Iron: Tuning the Properties of This Complex Material for Efficient Groundwater Remediation. Acc. Mater. Res. 2021;2(6)420–431. https://doi.org/10.1021/accountsmr.1c00037
crossref

40. Ling L, Pan B, Zhang W. Removal of selenium from water with nanoscale zero-valent iron: Mechanisms of intraparticle reduction of Se (IV). Water Res. 2015;71:274–281. https://doi.org/10.1016/j.watres.2015.01.002
crossref pmid

41. Ling L, Zhang W. Reactions of Nanoscale Zero-Valent Iron with Ni (II): Three-Dimensional Tomography of the "Hollow Out" Effect in a Single Nanoparticle. Environ. Sci. Technol. Lett. 2014;1(3)209–213. https://doi.org/10.1021/ez4002054
crossref

42. Farrell J, Kason M, Melitas N, Li T. Investigation of the long-term performance of zero-valent iron for reductive dechlorination of trichloroethylene. Environ. Sci. Technol. 2000;34(3)514–521. https://doi.org/10.1021/es990716y
crossref

43. Wang Y, Yu L, Wang R, Wang Y, Zhang X. A novel cellulose hydrogel coating with nanoscale Fe0 for Cr (VI) adsorption and reduction. Sci. Total. Environ. 2020;726:138625 https://doi.org/10.1016/j.scitotenv.2020.138625
crossref pmid

44. Li X, Elliott DW, Zhang W. Zero-Valent Iron Nanoparticles for Abatement of Environmental Pollutants: Materials and Engineering Aspects. Crit. Rev. Solid State Mater. Sci. 2006;31(4)111–122. https://doi.org/10.1080/10408430601057611
crossref

Fig. 1
XRD patterns of nZVI, S-nZVI, Ag/S-ZVI (S/Fe molar ratio=0.5).
/upload/thumbnails/eer-2022-379f1.gif
Fig. 2
SEM images of Ag/S-ZVI (S/Fe molar ratio=0.5; Fe/Ag molar ratio=40:1).
/upload/thumbnails/eer-2022-379f2.gif
Fig. 3
Different factors on the degradation rate of TCE: (a) materials; (b) one-order kinetic fitting model; (c) S/Fe molar ratio; (d) Ag+ addition; (e) initial pH; (f) initial TCE concentration and the kinetic fitting: (g) pseudo-first-order kinetics fitting, (h) pseudo-second-order kinetics fitting. (Dosing amount=0.5g/L; T =25°C; n=2)
/upload/thumbnails/eer-2022-379f3.gif
Fig. 4
SEM images of Ag/S-ZVI (a, c) and SEM images of the material after reaction with TCE solution (50 mg/L) (b, d). (S/Fe molar ratio=0.5, Fe/Ag molar ratio=20:1).
/upload/thumbnails/eer-2022-379f4.gif
Fig. 5
Fe elemental content and XPS spectra of Ag/S-ZVI materials before and after the reaction. (a) Fe 2p; (b) Fe valence content.
/upload/thumbnails/eer-2022-379f5.gif
Fig. 6
S elemental content and XPS spectra of Ag/S-ZVI materials before and after the reaction. (a) S 2p; (b) S valence content.
/upload/thumbnails/eer-2022-379f6.gif
Fig. 7
Ag elemental content and XPS spectra of Ag/S-ZVI materials before and after the reaction. (a) Ag 3d; (b) Ag valence content.
/upload/thumbnails/eer-2022-379f7.gif
Fig. 8
A proposed model on the reaction mechanism of TCE removal by Ag/S-ZVI.
/upload/thumbnails/eer-2022-379f8.gif
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