Article Cite This: Environ. Sci. Technol. 2018, 52, 14352−14361
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Mechanisms of Interaction between Persulfate and Soil Constituents: Activation, Free Radical Formation, Conversion, and Identification Guodong Fang,† Xiru Chen,†,§ Wenhui Wu,† Cun Liu,† Dionysios D. Dionysiou,‡ Tingting Fan,∥ Yujun Wang,*,† Changyin Zhu,† and Dongmei Zhou*,†
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†
Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, P.R. China ‡ Environmental Engineering and Science Program, Department of Chemical and Environmental Engineering (ChEE), University of Cincinnati, Cincinnati, Ohio 45221-0071, United States § University of Chinese Academy of Sciences, Beijing 100049, P.R. China ∥ Nanjing Institute of Environmental Science, Ministry of Environmental Protection of the People’s Republic of China, Nanjing 210042, P.R. China S Supporting Information *
ABSTRACT: Persulfate-based in situ chemical oxidation (ISCO) for soil remediation has received great attention in recent years. However, the mechanisms of interaction between persulfate (PS) and soil constituents are not fully understood. In this study, PS decomposition, activation, free radical formation and conversion processes in 10 different soils were examined. The results showed that soil organic matter (SOM) was the dominant factor affecting PS decomposition in soil, but Fe/Mn-oxides were mainly responsible for PS decomposition when SOM was removed. Electron paramagnetic resonance (EPR) spectroscopy analysis showed that sulfate radicals (SO 4 •−) and hydroxyl radicals (•OH) generated from PS decomposition subsequently react with SOM to produce alkyl-like radicals (R•), and this process is dependent on SOM content. R• and SO4•−/•OH radicals predominated in soil with high and low SOM, respectively, and all three radicals coexist in soil with medium SOM. Chemical probe analysis further identified the types of radicals, and R• can reductively degrade hexachloroethane in high SOM soil, while SO4•− and •OH oxidatively degrade phenol in low SOM soil. These findings provide valuable information for PS-ISCO, and new insight into the role of SOM in the remediation of contaminated soil.
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INTRODUCTION
nanodiamonds, and biochar have also been developed to activate PS for contaminant degradation.15−20 Among PS activation processes, base, iron, and chelated iron are the most commonly used activators in field applications.7 Watts and co-workers elucidated the mechanism of free radical formation and evolution in the base/PS system.7 Meanwhile, Liang and co-workers reported that PS activated by ferrous ions can efficiently degrade contaminants, and chelated iron substantially increases the activation efficiency and extends the usable pH range of this process.21,22 Base activation is suitable for degrading chlorinated solvents such as tetrachloroethylene, trichloroethylene, and dichloroethylene, while chelated iron activation processes can degrade aromatic hydrocarbons,
In situ chemical oxidation (ISCO) technology is a process of injecting or adding oxidants to the subsurface to remediate contaminated soil and groundwater, and it has been receiving increasing attention in recent years.1 Compared with conventional oxidants such as hydrogen peroxide (H2O2) and ozone (O3), persulfate (PS) is more stable and easier to transport in the subsurface environment, and more reactive with a wider range of pollutants than permanganate, which makes PS more suitable for the remediation of contaminated sites using ISCO processes.2 Consequently, PS-based advanced oxidation technologies have grown in popularity for the treatment of contaminants in both laboratory studies and ISCO field applications.3 The activation of persulfate (PS) by UV, heat, base, transition metals, and related metal oxides produces sulfate radicals (SO4•−) and hydroxyl radicals (•OH) that can degrade a wide range of contaminants.4−14 In addition to these conventional activation methods, some new metal-independent activators such as graphene, carbon nanotubes, graphitized © 2018 American Chemical Society
Received: Revised: Accepted: Published: 14352
August 24, 2018 November 7, 2018 November 14, 2018 November 14, 2018 DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology
Figure 1. Decomposition of persulfate (PS) in soils with different properties: (a) the observed reaction rate constant (kobs) of PS decomposition in different soils; (b) changes in pH of soil slurries during PS decomposition; and (c) correlation analysis between the content of soil organic matter (SOM) and the rate of PS decomposition. In reactions, the initial concentration of PS was 5.0 g/L in aqueous solution or 10 g/kg in soil, and the soil water ratio was 1:2 at 25 °C.
methyl-tert-butyl ether and 1,4-dioxane efficiently in field applications.23,24 When PS solution is injected or added to the subsurface, it is consumed by target contaminants as well as natural organic matter and other reduced species.25 In situ addition of PS solution to surface soil (depth 0.05). Due to the complexity of soil properties and the difficulty of isolating the effects of different variables, additional experiments were conducted to further investigate the sole SOM effect on PS decomposition in soil samples from different depths in a soil profile. Soil sampled at different horizons retained similar soil properties such as mineral contents, but the SOM content decreased successively from 35.5 to 6.36 g/ kg with increasing depth (Table S2). As shown in Figure S4, the rate of PS decomposition varied markedly among these soil samples, and exhibited a significant positive correlation with SOM content (R2 = 0.99, p = 0.0003). These combined results 14355
DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology
Figure 4. Degradation of the reductant probe HCA by PS in soil with high, medium and low SOM: (a) kinetics of HCA degradation by PS in high SOM soil (soil 6); (b) pseudo- first order rate constants of HCA degradation by PS in high, medium and low SOM soil; (c) kinetics of HCA degradation in the soil 6/PS system in the presence of alkyl radical quenchers such as O2 and Cr6+; and (d) kobs of HCA degradation in the presence of different quenchers. In reactions, the initial concentration of PS was 10 g/kg, and the soil water ratio was 1:2 at 25 °C, HCA and Cr6+ concentrations were 5.72 mg/kg and 2.0 mM, respectively.
consistent with •OH added to DMPO. Meanwhile, a six-line signal with a ratio intensity of 1:1:1:1:1:1 also accompanied the formation of DMPO-OH, indicating alkyl-like radical (R•) adducts to DMPO, based on the hyperfine splitting constants (aN = 15.3, aH = 22.9 G) obtained by simulation of the EPR spectrum (Figure S9). The hyperfine splitting constants of R• radicals are in the range of alkyl radicals (15 to17 G for aN, and 21 to 23 G for aH), hence R• radicals were defined as alkyl-like radicals according to a previous study.51 The intensity of DMPO-R increased markedly when the incubation time was extended from 1 to 5 days, which indicates that the concentration of R• increased with increasing reaction time. SO4•− is usually the source of other radicals generated from the decomposition of PS during PS activation processes, and the •OH is derived from the subsequent reactions of SO4•− with −OH or H2O.52 However, the EPR signal of SO4•− was insignificant at the beginning of the reaction. Two factors likely contributed to this phenomenon; (1) although SO4•− was generated from PS activation by Fe/Mn-oxides and SOM, it would be converted to •OH, and (2) the spin-trapping agent (DMPO) is more reactive with •OH than SO4•−, resulting in a higher intensity of the EPR signal of •OH compared with that of SO4•− at a similar concentration.30 Furthermore, broad peaks appeared at day 11, which were attributed to an increase in the amount of dissolved Mn species in the solution. The
correlated with kobs values. Regarding free Fe-oxides, a significant positive correlation was found between FeDCB content and kobs with a coefficient of 0.845 (p = 0.0001), while the content of other Fe-species ([Fe]total − [Fe]DCB) content was not significantly correlated with kobs (R2 = 0.03, p = 0.29; Figure S7). Similarly, in the case of free Mn-oxides, MnDCB content exhibited a significant positive correlation with kobs, with coefficients of 0.612 (p = 0.004), while the content of other Mn-species ([Mn]total − Mn]DCB) content was not significantly correlated with kobs (R2 = 0.14, p = 0.16; Figure S8). These results indicate that free Fe/Mn-oxides were the dominant Fe/Mn species responsible for PS decomposition in soil after removal of SOM. Free Radical Formation and Conversion during the Persulfate Decomposition. Decomposition of PS is usually accompanied by free radical formation and conversion.49,50 SOM is the dominant factor influencing PS decomposition in soil, as discussed above. Thus, free radical formation and conversion in three soil types (soil 3, low SOM, 2.2 g/kg; soil 4, medium SOM, 15 g/kg; soil 6, high SOM, 37.6 g/kg) were investigated using EPR spectroscopy coupled with DMPO as a spin-trapping agent to identify the dominant free radicals. As shown in Figure 3a, a significant EPR signal comprised of four lines with a ratio of 1:2:2:1 and hyperfine splitting constants of 14.8 G (aN = aH) was observed for soil 6 when PS was added, 14356
DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology
Figure 5. Degradation of the oxidative probe phenol by PS in soil with high (soil 6), medium (soil 4) and low SOM (soil 3): (a) kinetics of HCA degradation; (b) pseudo- first order rate constants of phenol degradation by PS in different soils; (c) influence of the SO4•−/•OH quencher ethanol on phenol degradation in the soil 3/PS system. In reactions, the initial concentration of PS was 5.0 g/L, and the soil water ratio was 1:2 at 25 °C. The phenol and ethanol concentrations were 5.72 and 941 mg/kg, respectively.
formation of SO4•−. However, the concentration of R• in soil 3 was much lower, and it did not change significantly over the incubation. Notably, the total •OH concentration decreased in all three soils, which can be ascribed to quenching reactions between SOM and •OH (Figure S10). The higher the SOM content in soil, the greater the quenching of •OH. These results indicate that •OH, SO4•−/•OH/R•, and R• radicals were the dominant reactive species in soils with low, medium, and high SOM, respectively, and their conversion processes were dependent on SOM content. Free Radical Identification with Chemical Probes. Chemical probes were employed to explore the free radical formation and conversion processes in the soil/PS system. Phenol and HCA were selected as probes for •OH/SO4•− and R•, respectively, because the reduction potential of R• ranges from −1.72 to −1.19 V, and they can reductively degrade compounds via electron transfer or nucleophilic addition.55 R• Radical Identification. Figure 4a shows that HCA was significantly degraded in soil 6 (high SOM)/PS system, with 71.2% of HCA decayed within 9 days. However, HCA degradation in soil without SOM or PS was insignificant. For soil 4 with medium SOM (Figure S11a), 43% of HCA was degraded by R • within 21 days, due to a lower R • concentration than soil 6, as revealed by EPR spectroscopy analysis (Figure 3). Additionally, only 14.8% of HCA was decayed in soil 3 (low SOM) within 21 days (Figure S11b). The kobs value of HCA degradation in soil 6 was 0.141 day−1, which was approximately four and ten times higher than in soil 4 and soil 3, respectively (Figure 4b). HCA degradation was insignificant in all soils after SOM was removed (Figure S11).
results indicate that the dissolution of Mn ion during PS catalytic reactions especially in high SOM soil would potentially contribute to PS decomposition. Similarly, for soil 4, which contained a medium level of SOM, DMPO-OH was dominant within 5 days, and DMPO-R was formed as the incubation time was prolonged to 9 and 14 days (Figure 3b). For soil 3, which contained a low amount of SOM, DMPOOH was the dominant reactive species within the 21 days (Figure 3c). To further elucidate the mechanism of free radical conversion in this process, the peak intensity of the R• radical was normalized with the peak intensity of DMPO-OH, and the corresponding ratio was determined as a function of incubation time (Figure 3d). For soil 6, the percentage of DMPO-R increased from 14.3% to 94.3% on 1 day, then decreased to 23.5% by day 9, which indicates that •OH was converted to R• via reactions with SOM or its degradation intermediates. Moreover, R• would be expected to transfer electrons to PS to yield SO4•− due to the intrinsic electron transfer properties of R•,53 as evidenced by the formation of a significant DMPO− SO4 signal on day 1 (Figure 3a). Furthermore, R• and SO4•− coexist, while the •OH signal was largely reduced with prolonged reaction time. This could be attributed to the fact that once •OH is formed, it may be quickly quenched by SOM or SOM degradation products to form R•, and R• further activates PS to form SO4•−. These results also imply that SO4•− is less reactive toward SOM, and provides direct evidence suggesting that SO4•− is more stable than •OH in the subsurface environment.54 Similar results were also observed for soil 4 (Figure 3d); the ratio of DMPO-R to DMPO-OH peaked on day 9, then decayed by day 14, accompanying the 14357
DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology Scheme 1. Proposed Pathways of Persulfate Activation, Free Radical Formation, and Conversion in Soil
These results suggest that R• radicals mainly accounted for the reductive degradation of HCA. Free radical quenching studies were performed to further investigate the reductive degradation of HCA by R•, and both O2 and Cr6+ were selected as electron quenchers.56,57 Figure 4c shows that HCA degradation was greatly inhibited by purging O2 or adding Cr6+, and the corresponding kobs value decreased from 0.141 to 0.064 and 0.011 day−1, respectively. By contrast, HCA degradation was markedly enhanced by purging N2, with the corresponding kobs increasing from 0.141 to 0.236 day−1, since purging N2 eliminated the quenching reaction between R• and O2. Furthermore, purging with O2 or N2 and introducing Cr6+ had an insignificant effect on HCA degradation by PS in soil 6 without SOM (Figure S12). These results indicate that R• was the dominant radical formed in soil 6 with high SOM, consistent with the EPR spectroscopy results. Identification of •OH/SO4•− Radical. Figure 5a shows that 13.1%, 49.3%, and 64.9% of phenol was degraded within 21 days in soil 6 (high SOM), 4 (medium SOM), and 3 (low SOM), respectively. After removal of SOM, 69.8%, 49.4%, and 39.2% of phenol was degraded in soil 6, soil 4, and soil 3 (Figure S13a), respectively, which was ascribed to PS activation by Fe/Mn minerals to form SO4•− and •OH. The adsorption only led to less than 10% of phenol decay in these three soils (Figure S13b). Additionally, the kobs values for phenol degradation in soil 6 and soil 4 were 0.013 and 0.026 day−1, respectively (Figures 5b and S13c), significantly lower than in the equivalent soils without SOM (0.069 and 0.043 day−1). These results indicate that high SOM competes with phenol for the consumption of SO4•− and •OH, whereas R• radicals formed exhibit limited reactivity toward phenol, hence only a small amount of phenol was degraded in soil 6. However, the kobs value of phenol degradation in soil 3 (0.059 day−1) was markedly higher than in soil 3 without SOM (0.036 day−1). The likely reason is that PS was also activated by R• to produce SO4•−, as discussion in the above EPR results (Figure 3). SOM could enhance phenol degradation when the rate of SO4•− produced from PS activated by R• was greater than the rate of SO4•− consumed by SOM. These result were consistent with those of a previous study showing that quinones exhibit dual effects (enhancement or inhibition) regarding contaminant degradation by persulfate.30 Ethanol was used as a quencher to identify phenol oxidation by SO4•− and •OH. Purging with both N2 and O2 had an insignificant effect on phenol degradation (p > 0.05), while
ethanol greatly inhibited phenol degradation in soil 3, with the degradation efficiency decreasing dramatically from 65.1% to 15.3% (Figure 5c). These results indicate that SO4•− and •OH were the dominant reactive species involved in phenol degradation, while R• contributed minimally to phenol degradation directly in soils with low SOM. Some previous studies reported that nonradical and singlet oxygen (1O2) pathways contribute to pollutant degradation in PS activation processes.58−60 However, these processes would not be involved in the soil/PS system, as evidenced by radical quenching (Figure 5) and EPR experiments (e.g.,1O2 was not detected, data not shown). Proposed Pathways of PS Activation, Free Radical Formation, and Conversion in Soil. On the basis of the above results, PS activation, free radical formation, and conversion pathways are proposed in Scheme 1. Both SOM and free Fe/Mn-oxides account for PS activation and the production of SO4•− and •OH, but SOM was the dominant factor contributing to PS decomposition in ISCO when PS was added to soil. When SOM was largely consumed, Fe/Mnoxides were the dominant factor for PS activation. Two processes are involved in the SOM-PS interaction system. First, SOM is composed of various carbon functional moieties including alkyl, alkoxy, aromatic, and carbonyl groups.61,62 Phenol or quinone-type structures in SOM can induce the formation of semiquinone-like radicals that activate PS to produce SO4•− and •OH via single electron transfer processes.30 Meanwhile, SOM is degraded to produce lower molecular weight alkyl fragment (RH) products.63 Second, RH reacts with •OH or SO4•− via H-abstraction processes to yield R•, and R• further activates PS to produce SO4•− via electron transfer with second order reaction rate constants of ∼1.5 × 105 M−1 s−1.53 Importantly, free radical formation and conversion during SOM-PS interactions are dependent on SOM content. In soil with relatively high SOM, both SO4•− and •OH are eventually converted to R• radicals via reactions with SOM, and R• is the dominant reactive species accounting for the reductive degradation of probes. In soil with medium SOM, SO4•−, •OH, and R• coexist, but in soil with low SOM, SO4•−, and •OH are mainly responsible for oxidation of probes. Other Considerations for Free Radical Conversion in the Soil Environment. Alkyl-like Radicals (R•). R• radicals were indeed formed from the soil/PS system, as detected by EPR spectroscopy. However, the exact structures and types of R• could not be identified due to limitations of the EPR 14358
DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology technique, since different R• species such as methyl, ethyl, propyl, and other alkyl-containing macromolecule radicals display similar hyperfine splitting constants when DMPO is employed as a spin-trapping agent.51 Therefore, R• represents a mixture of carbon-centered radicals, and the composition is more complex than any single alkyl radical such as •CH3 or •CH3CH2 due to the structural complexity of SOM. Other sensitive methods must be developed to characterize the exact structures of the alkyl radicals formed in the complex soil environment. Alkylperoxyl Radicals (RO2•). When R• radicals are formed, they react with oxygen (O2) to produce RO2•, then induce free radical chain reactions (eqs 1−3).64 RO2• with reduction potentials of ∼0.71−0.94 V,65 and can further oxidize RH via H-abstraction to regenerate R• (eq 2) .64 RO2• also decays with quenching reactions according to eq 3. R• + O2 → RO•2
(1)
RO•2 + RH → RO2 H+R•
(2)
2RO•2 → RO4 R → products
(3)
transport distance, and remediation efficiency of PS when following delivery or diffusion into various soil layers during ISCO treatments. More importantly, our study revealed that SOM plays an important role in free radical formation, conversion, and distribution after PS addition, which directly dominates the transformation of many contaminants. In soil layers with low SOM content, Fe/Mn-oxide-rich mineral phases contribute to PS decomposition, and influence the persistency of PS in soil matrices. In particularly, organic matter is thought to quench free radicals, and thus inhibit the degradation of target contaminants in the soil environment. However, the results of the present study showed that oxidative radicals such as SO4•− and • OH can be converted to R• via reaction with organic matter, which indicates great potential for the reductive degradation of contaminants such as halogen compounds. Therefore, the application of PS in SOM-rich soils, especially surface soils, favors the reductive degradation of contaminants, while PS contributes to the oxidative degradation of contaminants in subsurface soils with low SOM content. In summary, the results confirm the qualitative dependency of PS decomposition and contaminant degradation pathways on the distribution of SOM and Fe/Mn content along soil profiles.
The rate constant for the trapping of RO2• by DMPO was reported to be much lower than that of •OH and R•, and the DMPO-RO2 adduct is unstable.66 Consequently, RO2• was not detected in soil/PS in the present study. Additionally, R• may react preferentially with S2O82− instead of O2 to produce SO4•− in the soil/PS system. Superoxide Anion Radical (O2•−). Reductive O2•− is generated via the acceptance of an electron by O2 or free radical chain reactions during persulfate activation processes,5,67 but this species was not detected by EPR spectroscopy in the present study. Additionally, the concentration of O2•− was significantly lower than that of R•, and the reactivity of O2•− (∼−0.33 V) toward HCA was lower than that of R• (∼−1.72 to −1.19 V). Therefore, we hypothesized that O2•− makes only a limited contribution to HCA degradation, and this was supported by the observation that an increase in dissolved oxygen (DO) concentration inhibited HCA degradation in the soil 6/PS system (Figure 4c), even though the formation of O2•− was positively correlated with DO concentration. Environmental Implications. PS-based ISCO has been extensively used for the remediation of contaminated subsurface soil and groundwater, and PS-based technologies have recently been extended to surface soil in the unsaturated zone.68,69 The results of present study showed that the interactions between soil components such as SOM and Fe/ Mn-oxide with PS not only influence the decomposition of PS, but also affect free radical formation, conversion, and distribution. This demonstrates the potential for alternative ISCO strategies and improves our understanding of the environmental behavior of PS in the broader soil layers. The long residence time and large diffusion radius of PS are key to its effective remediation of contaminants spread over broader soil horizons. The recent application of PS coupled with in situ stirring technologies for treating surface soil demonstrates its promise for general remediation not limited to the surbsurface.70 We therefore characterized interactions between PS and soil components in samples from different soils and soil horizons. The results showed that SOM dominates PS decomposition in soil, which indicates that interactions between PS and SOM control the persistency,
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ASSOCIATED CONTENT
S Supporting Information *
. This material is available free of charge via the Internet at http://pubs.acs.org/. The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.8b04766. The properties of soils used in this study and the rates of PS decomposition in different soils, Tables S1, S2, and S3; and the kinetics of PS decomposition, correlation analysis, EPR simulation, and HCA and phenol degradation, Figures S1−S13 (PDF)
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AUTHOR INFORMATION
Corresponding Authors
*Phone: 86-25-86881180; fax: 86-25-86881180; e-mail:
[email protected] (Y.J.W.). *Phone: 86-25-86881180; fax: 86-25-86881180; e-mail:
[email protected] (D.M.Z.). ORCID
Guodong Fang: 0000-0002-3837-6279 Dionysios D. Dionysiou: 0000-0002-6974-9197 Yujun Wang: 0000-0002-0921-0122 Dongmei Zhou: 0000-0002-7917-7954 Notes
The authors declare no competing financial interest.
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ACKNOWLEDGMENTS
This work was supported by grants from the National Key Research and Development Program of China (2017YFA0207001, 2016YFD0800204), the Natural Science Foundation of Jiangsu Province of China (BK20170050), the National Natural Science Foundation of China (41671478), the Youth Innovation Promotion Association of CAS (2014270), and the 135 Program of Institute of Soil Science (ISSASIP1660). 14359
DOI: 10.1021/acs.est.8b04766 Environ. Sci. Technol. 2018, 52, 14352−14361
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Environmental Science & Technology
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