Biomimetic degradation of perfluorinated acids by vitamin B12 with nano-zero-valent iron/nickel bimetal: effects of their self-structure and coexisting substances

Highlights

	● Degradation of long-chain PFAs is better than short-chain in VB12 + nFe0/Ni0 systems.
	● PFSAs are more susceptible to defluorination and removal than PFCAs in this system.
	● Degradation products of some PFAs were identified and possible pathways were proposed.
	● The system has good anti-interference ability to common natural water components.

Abstract

Perfluorinated acids (PFAs) are a new class of persistent organic pollutants that are difficult to defluorinate or remove. The reductive degradation of various representative PFAs in a biomimetic system composed of vitamin B₁₂ (VB₁₂) as a catalyst and nano-zero-valent iron-nickel bimetal (nFe⁰/Ni⁰) as a reductant was investigated in this study. The effects of the self-structures of PFAs and the coexisting substances in natural water were also discussed. The results indicated that the defluorination and removal rates of PFAs were highly dependent on the length and terminal functional groups of the perfluorocarbon chain. Only Perfluorocarboxylates with C > 11 and Perfluorosulfonates with C > 6 were significantly degraded. Based on the analysis of the degradation products of perfluorobutanesulfonate (PFBS), perfluorohexanesulfonate (PFHxS), prefluorooctanesulfonate (PFOS), and 2-perfluoroctyl ethanol (8:2 FTOH), hydrolysis followed by the scission of C–S or C–C connecting the terminal functional groups was the dominant degradation pathway of long-chain PFAs instead of cleavage of C–C in the perfluorocarbon chain. The perfluorocarbon chain length affects the product type. It is speculated that the high bond dissociation energies of C–F bonds in short-chain PFAs hinder the occurrence of the decarboxylation-hydroxylation-elimination-hydrolysis (DHEH) pathway and make the addition of (–CF₂–)_n dominant. Meanwhile, the inhibition of SO₄^2– removal by PFOS was significant, whereas humic acid, Cl^–, and dissolved oxygen had only a slight influence. Overall, this study provides new insights on the degradation of PFAs containing multiple structures and highlights the impact of the self-structure on PFAs removal.

Graphical abstract

Keywords

Perfluorinated compounds / Vitamin B₁₂ / nFe⁰/Ni⁰ / Biomimetic reduction

1 Introduction

Perfluorinated acids (PFAs) have been used extensively since they were first synthesized in the late 1940s. Owing to their oleophobic and hydrophobic properties, as well as their resistance to acid and high temperatures, they are used in various industrial products (e.g., firefighting foams, food packaging, pharmaceuticals, leather, synthetic detergents, and pesticides) (Prevedouros et al., 2006; Paul et al., 2009). PFAs exhibit toxicity in multiple aspects such as reproduction, development, and neurological and immunological processes, posing a great threat to human health (Rahman et al., 2014; Wang et al., 2017; Qiu et al., 2020; Wan et al., 2020; Liang et al., 2022). Moreover, PFAs endanger the environmental and ecological security. Due to their ability to bioaccumulate and migrate long distances, PFAs can be detected not only in a variety of environmental media, including drinking water, air, dust, and food but also in human blood and wildlife (Qiu et al., 2010; Domingo, 2012; Mamsen et al., 2017; Yang et al., 2023). Furthermore, PFAs are widespread in remote Arctic environments through rainfall and snowfall (Butt et al., 2010; Taniyasu et al., 2013). Toxicity research on classical terrestrial biota has shown the damage caused by PFAs to the soil environment from both macro-indicator and molecular perspectives (Cai et al., 2021). Due to the high electronegativity and small atomic radius of the fluorine atom, the C–F bond is one of the highest bond energies in nature (Bentel et al., 2019). Therefore, PFAs have extremely high thermal and chemical stabilities, making them very difficult to degrade through traditional water treatment processes (Liou et al., 2010; Krafft and Riess, 2015). The manufacturing, monitoring, and remediation of PFAs have received considerable attention. In particular, it is necessary to develop in situ application technologies because of the ineffectiveness of traditional methods.

Many countries have introduced regulations regarding restrictions on PFAs. For example, prefluorooctanesulfonate (PFOS) and its salts were listed in the Stockholm Convention on Persistent Organic Pollutants in 2009 (UNEP, 2009). Limits for perfluorooctanoic acid (PFOA) and PFOS have been set in multiple US states, including California, Massachusetts, Michigan, and others (Andrews and Naidenko, 2020). The US Environmental Protection Agency has established Maximum Contamination Levels (MCLs) for six per- and polyfluoroalkyl substances (PFASs) in drinking water, with the enforceable level for PFOA and PFOS being 4.0 ng/L (EPA, 2024). These regulations have triggered significant efforts to develop remediation technologies of PFAs. Physical separation (e.g., activated carbon, zeolite, and anaerobic sludge) has a good removal rate of PFAs; however, wastes containing PFAs must be treated through subsequent processes to destroy them, which may cause secondary pollution (Arvaniti et al., 2014; Li et al., 2019; Saeidi et al., 2020; Xie et al., 2020). Chemical methods, such as ultrasonic oxidation, photochemical, and electrochemical methods work well for the defluorination of PFAs (Niu et al., 2016; Liu et al., 2020a; Tan et al., 2023). However, these advanced oxidation or reduction techniques are usually performed under harsh conditions, which leads to large consumption of energy or difficulty in in situ remediation (Trojanowicz et al., 2018; Wang et al., 2021). Another way to deal with PFAs is through biodegradation, which involves defluorination and removal through enzymes in microorganisms (Hua et al., 2022; Kuok Ho and Kristanti, 2022). Biodegradation is an economical and environmentally friendly technology that consumes less energy than chemical methods. This is a promising method for the in situ remediation of water-containing PFAs (Leung et al., 2022). However, biodegradation requires the cultivation of specific microorganisms, and effective defluorination is difficult. In addition to biodegradation, biomimetic degradation has outstanding prospects for application in situ remediation (Leung et al., 2022).

Vitamin B₁₂ (VB₁₂) is a transition metal coenzyme produced by anaerobic soil microorganisms that is used as a catalyst for typical biomimetic degradation. Previous studies have shown high catalytic activity for most halogenated compounds (Pratt and van der Donk, 2006; Heckel and Elsner, 2022), and high efficiency in defluorination (Im et al., 2014). Ochoa-Herrera et al. (2008). first constructed a system that used the VB₁₂ as catalyst and Ti(III)-citrate as the reductant to degrade PFOS. A previous study found that VB₁₂ with Ti(III)-citrate and Cu⁰ generated extraordinarily reactive radicals for PFOA defluorination and removal (Lee et al., 2017). Park et al. (2017) explored the effects of different alternative reductants, such as nZn⁰, nFe⁰, and Pd/nFe⁰ on linear and branched PFOS. These studies showed that the system of VB₁₂ with various reductants has significant effects on PFAs. It was found that the bimetallic material effectively enhanced the reactivity of VB₁₂ compared with various reducing agents. Benefiting from the strong synergistic effect between nickel and iron, nickel/zero-valent iron can accelerate the dechlorination of common halogenated pollutants, such as trichloroethylene and polychlorinated biphenyls (Han et al., 2018; Wu et al., 2023). Our research group also verified the effect of adding different transition metals to composite bimetallic materials on the degradation of PFOS. Compared with Fe/Cu and Fe/Pd, Fe/Ni had a higher defluorination degradation effect on PFOS. These studies focused on determining the optimal reductant and developing a technology aimed only at degrading PFOA and/or PFOS. However, the PFAs used in industry are large and contain a variety of structures. Furthermore, because of regulations and restrictions on the use of typical PFAs, many industries have shifted toward using newer replacement PFAs, such as hexafluoropropylene oxide dimer acid and perfluorohexanesulfonic acid (PFHxS), while the understanding of the environmental fate and remediation technologies of these alternatives is still limited (Ateia et al., 2019). The chemical properties of these substitutes are similar to those of typical PFAs; however, they may not degrade well in systems where PFOA or PFOS are effectively degraded, due to differences in their chemical structure. For example, a UV-sulfite system was used to degrade commercial aqueous film-forming foam (AFFF), a complex PFASs mixture consisting of seven perfluoroalkyl sulfonic acids (PFSAs, 3C–7C), five perfluoroalkyl carboxylic acids (PFCAs, 4C–8C), and three fluorotelomer sulfonic acids (FTSs, 6C, 8C, 10C), and showed that the reactivity varied widely among PFASs (Tenorio et al., 2020). This indicateds that the self-structures of different PFAs affected their degradation rates in the same system. Studies have reported the degradation of branched PFAs with different structures, reflecting the structure-reactivity relationships within branched PFAs in VB₁₂ and Ti(III)-citrate system (Liu et al., 2018). However, it is not sufficient to assess the impact of branched structures alone. Accordingly, it is important to comprehensively investigate the structure-reactivity of PFAs in biomimetic systems.

This study explored the defluorination and removal rates of PFAs with different chemical structures and coexisting substances such as natural organic matter (NOM), dissolved oxygen (DO), Cl^– and SO₄^2– under a biomimetic system composed of VB₁₂ and nFe⁰/Ni⁰. The purpose of this study was to determine the structure-reactivity relationships with PFAs and analyze the influence of different coexisting substances by comparing experimental results, target pollutants with high removal rates, and give full play to the advantages of the system to provide a feasible solution for the in situ remediation of water sources polluted by these PFAs.

2 Methods and materials

2.1 Standards and chemicals

Perfluorobutanoate (PFBA, 98%), perfluoroheptanoate (PFHpA, 97%), PFOA (95%), perfluorononanoate (PFNA, 97%), perfluorododecanoate (PFDoA, 95%), perfluorobutanesulfonate (PFBS, 98%), perfluorohexanesulfonate (PFHxS, 98%), 2-perfluoroctyl ethanol (8:2 FTOH, 98%) and PFOS (95%) were obtained from Sigma-Aldrich (Milwaukee, WI, USA). HPLC grade methanol (≥ 99.9%) and ammonium acetate (≥ 99.0%) were acquired from Sigma-Aldrich and Aladdin, respectively. Analytical grade formic acid (≥ 95.0%) was obtained from Sigma-Aldrich. Envi-C18 SPE cartridges (3 mL, 500 mg) and weak anion-exchange (WAX) cartridges (3 mL, 200 mg) were purchased from Supelco (Bellefonate, PA, USA) and CNW, respectively. Except for polyethylene glycol, which was chemically graded, all other inorganic compounds were of analytical grade and were obtained from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China), unless otherwise specified. Milli-Q water (Milli-Q® Reference, Merck Millipore, Germany) was used in all experiments.

0.5 wt% nFe⁰/Ni⁰ was prepared according to the following method, with slight modifications based on a previous study (Yang et al., 2013). 0.5 wt%, indicating that Ni accounted for 0.5% of the total weight of the bimetal. FeSO₄, NaBH₄ and NiCl₂ solutions were prepared using oxygen-free deionized water and sealed with a sealing film. Polyethylene glycol (1.0 g) was then added to the FeSO₄ solution, and stirred until the solid polyethylene glycol completely dissolved; the NaBH₄ solution was added dropwise by peristaltic pump. After vigorous stirring for 30 min, a black suspension was obtained. To replace Fe on the surface of the material, NiCl₂ solution was added dropwise. All the operations were performed under a nitrogen atmosphere. The resulting 0.5 wt% nFe⁰/Ni⁰ was washed using oxygen-free Milli-Q water and absolute ethanol. 0.5 wt% nFe⁰/Ni⁰ was then placed in a vacuum dryer for 12 h at 60 °C.

2.2 Reductive decomposition

20 mL headspace glass vial was used as the reaction container and wrapped with aluminum foil to prevent the decomposition of VB₁₂ by light. All treatments were performed under anaerobic conditions, and the solution was blown with N₂ gas (99.999%) for 15 min prior to use. PFAs stock solution (500 μmol/L), VB₁₂ stock solution (2 mmol/L) and Milli-Q water were mixed. The 0.030 g 0.5 wt% nFe⁰/Ni⁰ was added to the headspace vial. The final volume was 10mL. The initial concentrations of PFAs, VB₁₂ and 0.5 wt% nFe⁰/Ni⁰ were 200 μmol/L, 200 μmol/L and 3.0 g/L, respectively. The solution pH was adjusted to 9.5 with NaCO₃ solution. Sealing film (Parafilm M®) was used to seal the vials. All vials were incubated at 60 °C in an air shaker or water bath shaker at 150 r/min.

2.3 Sample pretreatment

Sample pretreatment was based on a previous study with some modifications for accurate detection (Text S1) (Li et al., 2020). The samples were then subjected to solid-phase extraction (SPE) to remove interfering compounds and salts. Specific SPE operations were adjusted based on a previous study (Text S2) (Higgins et al., 2005). Samples containing PFBS and PFBA were extracted using WAX cartridges (3 mL, 200 mg), whereas other samples were extracted using Envi-C18 SPE cartridges (3mL, 500 mg).

2.4 Instrument analysis

The concentration of fluoride was determined using a fluoride ion selective electrode. PFAs were qualitatively and quantitatively analyzed using high performance liquid chromatography-tandem mass spectrometry (HPLC-MS/MS). The operating parameters for the instrument are provided in the Supporting Information. Ultra-performance liquid chromatography-quadrupole time-of-flight mass spectrometry (UPLC-QTOF) was used to analyze PFAs and their degradation products. More information is available in the Supplementary material.

2.5 Evaluation standards

The defluorination and removal effects of PFAs were measured using the defluorination and removal rates, which were defined as follows:

$\mathrm{D}\mathrm{e}\mathrm{f}\mathrm{l}\mathrm{u}\mathrm{o}\mathrm{r}\mathrm{i}\mathrm{n}\mathrm{a}\mathrm{t}\mathrm{i}\mathrm{o}\mathrm{n}\mathrm{r}\mathrm{a}\mathrm{t}\mathrm{e}=\frac{C_F-C}{A\times C_0}\times 100\mathrm{\%},$

$\mathrm{R}\mathrm{e}\mathrm{m}\mathrm{o}\mathrm{v}\mathrm{a}\mathrm{l}\mathrm{r}\mathrm{a}\mathrm{t}\mathrm{e}=\frac{C_0-C_t}{C_0}\times 100\mathrm{\%},$

where $C_F$ and C was the concentration of fluoride ions at time t and at the start of the reaction, μmol/L. $C_0$ and $C_t$ were the PFAs concentrations at initial and time t, μmol/L. A was the number of fluorine atoms in each PFAs.

2.6 QA/QC

Materials containing Teflon were not used in the experiment to reduce the instrumental background of the target analytes. The test tubes, centrifuge tubes, and volumetric flasks used in the experiments are made of polypropylene. Tubing made of Teflon in the instruments was replaced by polyetheretherketone or stainless-steel tubing. To reduce the effects of contaminants in the tubing, blank water samples (Milli-Q water only) were run three times before injection. The order of sample injection was from low concentration to high, and three blank water samples were collected to remove residual pollutants after running the high-concentration samples. A blank water sample was analyzed after five samples. After the injection, the syringe was rinsed with methanol. The standard curve contained at least five valid points, and the linear correlation coefficient (R² > 0.99). Meanwhile, a standard sample with a certain concentration was run every 10 samples to monitor contamination. When the peak areas of the standard changed by > 30%, the standard curve was recalibrated.

Several controls are carried out simultaneously with the decomposition treatment, referred to as VB₁₂ control (VB₁₂ + PFAs), 0.5 wt% nFe⁰/Ni⁰ control (0.5 wt% nFe⁰/Ni⁰ + PFAs), and container adsorption control (PFAs only), respectively.

3 Results and discussion

3.1 Effect of the length of perfluorocarbon chain

3.1.1 Perfluorocarboxylates (PFCAs)

Biomimetic degradation of representative PFCAs with different perfluorocarbon chain lengths, including PFBA (C3), PFHpA (C6), PFOA (C7), PFNA (C8), and PFDoA (C11), were investigated. As shown in Fig.1(a), the defluorination rates of all the adsorption controls were less than 0.2%, indicating that the PFCAs were not defluorinated in the absence of VB₁₂ and 0.5 wt% nFe⁰/Ni⁰. The defluorination rates of PFCAs with perfluorocarbon chain lengths of less than 11 were extremely low (< 0.5%) and were not significantly different from those of the adsorption controls (p > 0.05). However, the defluorination rate of PFDoA was higher than that of the other PFCAs, indicating that only PFDoA was defluorinated in this system. Similar phenomena were also reported in a UV/electrochemical system, which showed that the F^– concentration of PFNA was 1.12 mg/L, and the defluorination rate was significantly better than that of PFOA and PFHpA under UV/-2V conditions (Rao et al., 2020). The concentration of F^– in PFDoA decomposition treatment reached 168.42 μmol/L in 10 d and the defluorination rate was 3.66%, which was approximately 3.5 times that of 2 h (Fig.1(b)). The increase in the defluorination rate in the 24 h period was remarkably faster than that in the later period. The 24-h defluorination rate accounted for 75% of the final defluorination rate.

Fig.1 Time profiles of defluorination rate of adsorption controls (a), and decomposition treatments (b), removal rate of adsorption controls (c), and decomposition treatments (d) for PFCAs. The experiment conditions: [PFCAs]0 = 200 μmol/L, [VB12]0 = 200 μmol/L, [0.5 wt% nFe0/Ni0]0 = 3.0 g/L, pHi = 9.5, and t = 60 °C.

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As presented in Fig.1(c), the removal rates in the adsorption controls fluctuated only slightly in the initial period and tended to stabilize in the later period. These fluctuations are attributed to the adsorption of PFCAs onto the headspace glass vial wall. To eliminate the influence of adsorption, the removal rates of the decomposition treatments were deducted from those of the adsorption controls. The removal rates of all PFCAs first increased from 2 h and then decreased after 72 h. Although the effects of adsorption were removed, the 72 h removal rates of PFNA and PFHpA were still up to 8.94% and 8.74%, respectively. The reason for this phenomenon may be that a portion of the PFNA and PFHpA structures changed through C–C cleavage in the perfluorocarbon chain, but F^– was not generated (Sun et al., 2021). However, a significant removal rate of PFDoA was observed. After 10 days of incubation, the removal rate of PFDoA consistently increased from 0.71% at 2 h to 58.5% (Fig.1(d)). The reaction of PFDoA was almost complete after 24 h, reaching 63.0%. After 72 h, the highest removal rate of 65.5% was achieved. This rate exhibits a downward trend during the later period. This downward trend can be attributed to the recombination of the incompletely mineralized carbon-chain-shortened intermediates. Combined with previous studies, the possible degradation pathway of PFCAs is considered to be a shortening of the perfluorocarbon chains and H/F exchange (Qu et al., 2010; Gu et al., 2016; Cui et al., 2020). It was found that PFOS breaks at multiple sites on the perfluorocarbon chain during decomposition and from short-chain PFAs intermediates in a system composed of VB₁₂ and Ti(III)-citrate. Due to the recombination of the intermediates, PFDoA, with a longer carbon chain length than PFOS, was generated (Li et al., 2020). As shown in Fig.1(b) and Fig.1(d), the defluorination rate of PFDoA is significantly lower than the removal rate of PFDoA. This phenomenon is attributed to the incomplete mineralization of PFDoA, which produces fluorine-containing intermediates and degradation products. This indicates that the system still needs to be optimized to improve the defluorination rate.

The experimental results showed that the reaction of the PFCAs was highly dependent on the length of the perfluorocarbon chain. In this system, PFDoA (C11) exhibited more significant defluorination and removal than the other PFCAs. According to our group, PFTeDA (C13) also exhibited a significant removal rate in a similar biomimetic system (data not shown). Therefore, only long-chain PFCAs (C > 11) were significantly removed and defluorinated by the treatment with VB₁₂ and 0.5 wt% nFe⁰/Ni⁰.

3.1.2 Perfluorosulfonates (PFSAs)

The effects of the perfluorocarbon chain length on the degradation of PFSAs including PFBS (C4), PFHxS (C6), and PFOS (C8), were also investigated, and the results are presented in Fig.2. As illustrated in Fig.2(a), no defluorination occurred in any adsorption controls, and the highest defluorination rate was observed at the end of the reaction (0.26%). The defluorination rate of PFBS, a representative short-chain PFSAs, did not change significantly. Except for the 144-h defluorination rate of PFHxS, the defluorination of PFHxS and PFOS maintained a consistent growth trend. The defluorination rate of PFHxS increased 2.8 times from 3.39% at 2h to 9.45% at 240 h. The defluorination rate of PFOS on 10 d was 12.5%, which was 2.3 times of 2-h rate. Furthermore, the defluorination rate of PFOS in the entire reaction was higher than that of PFHxS, and the final defluorination rate was 1.32 times higher than that of PFHxS.

Fig.2 Time profiles of defluorination rate of adsorption controls and decomposition treatments (a), removal rate of adsorption controls (b), and decomposition treatments (c) for PFSAs. The experiment conditions: [PFSAs]0 = 200 μmol/L, [VB12]0 = 200 μmol/L, [0.5 wt% nFe0/Ni0]0 = 3.0 g/L, pHi = 9.5, and t = 60 °C.

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As shown in Fig.2(b), only the removal rate of PFBS adsorption control fluctuated irregularly. Large fluctuations occurred during the first 12 h, with negative removal rates at 2, 6, and 12 h. After 144 h, the removal rate increased sharply to 30.8%. This may have been due to the intensified molecular motion of short-chain PFSAs at high temperatures, leading to the formation of other substances. As illustrated in Fig.2(c), the 10-d removal rates of the three PFSAs were 11.4%, 34.0%, and 46.1%, respectively. Overall, all the removal rates showed an upward trend. In particular, PFOS maintained a relatively stable rate after 144 h, whereas PFBS and PFHxS continued to increase during the same period, indicating that after 10 d of treatment in this biomimetic system, PFBS and PFHxS still had the potential to be removed. The PFSAs removal and defluorination rates increased linearly with increasing perfluorocarbon chain length. Moreover, the defluorination and removal trends of PFSAs were almost synchronized. These results agree with those reported by Park et al. (2009), even in systems utilizing hydrated electrons for reductive defluorination. This suggests that the biomimetic system can remove and defluorinate medium- and short-chain PFSAs (C > 6).

According to the above results, the defluorination and removal rates of both PFCAs and PFSAs were highly correlated with the length of the perfluorocarbon chains in this system. The strong correlation between the removal and defluorination rates of PFAs and their self-structures may be determined from the bond dissociation energies (BDEs) of the C–F bond. The BDEs of C–F bonds are considered to be related to their positions on the perfluorocarbon chain. For example, the structure of PFBS contains only one low BDE, –CF₂– (BDE < 107 kcal/mol), whereas that of PFHxS contains three. This suggests that PFHxS is more susceptible to degradation via the decomposition of –CF₂– (Bentel et al., 2019). In addition, when the perfluorocarbon chain length increased, the proportion of –CF₂– with a low BDE in the PFAs increased. However, it is doubtful whether PFCAs with C > 11 have a remarkable removal rate. This shows that the BDEs of C–F bonds are not a decisive factor in the defluorination and removal rates of PFAs. Furthermore, VB₁₂ has been shown to be a nucleophile in previous studies and can be reduced in a system to attack halocarbon bonds (Dror et al., 2012; Lapeyrouse et al., 2019). Long-chain PFAs have more attack sites, so they may exhibit outstanding removal and defluorination rates.

3.2 Effect of terminal functional groups

The effect of the terminal functional groups was investigated by comparing PFCAs and PFSAs with the same perfluorocarbon chain length. A comparison of the defluorination and the removal rates at 72 h is shown in Fig.3. As illustrated in Fig.3(a), the defluorination rate of PFHxS was 29 times that of PFHpA (C6), and the defluorination rate of PFOS was as high as 122 times that of PFNA (C8). According to Fig.3(b), the removal rate of PFHxS was approximately three times higher than that of PFHpA, nearly 3 times. After 72 h, the removal rate of PFHpA fluctuated slightly, whereas that of PFHxS maintained an upward trend. In addition, the degradation rate of PFOS after 72 h was 3.5 times that of PFNA. In general, if the perfluorocarbon chain length is six or eight, the removal and defluorination rates of PFSAs with sulfonic acid functional groups are superior to those of PFCAs with carboxylic acid functional groups. In a reduction system using Mg-amino clay-coated nanoscale zero-valent iron, the removal rate of PFOS was higher than that of PFNA (Arvaniti et al., 2015). However, this differs from the results for a system composed of UV and sulfite with the generation of hydrated electrons, in which PFSAs with the same chain length showed worse efficiency than PFCAs (Bentel et al., 2019; Tenorio et al., 2020). This may be attributed to the different types and amounts of reductants and the dosages of PFAs. It is speculated that the reason for better removal rates of PFSAs in the system is that PFSAs are more likely to change their structure upon the cleavage of multiple bonds in the perfluorocarbon chain. Because of the single-electron transfer of VB₁₂, various reactions, such as C–C scission in the center of the perfluorocarbon chain, desulfonation, or H/F exchange, may occur (Sun et al., 2021). In addition, the C–S bonds in PFSAs are easier to remove because of their lower bond energies and BDEs (Gu et al., 2016). The proposed degradation pathways are discussed below.

Fig.3 The (a) defluorination and (b) removal rates of PFCAs and PFSAs with the same chain length at 72 h.

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3.3 Degradation products analysis

Under the condition that the initial concentration of PFAs and VB₁₂ was 20.0 μmol/L, and the dosage of 0.5 wt% nFe⁰/Ni⁰ was 7.0 g/L, the samples to be tested were incubated for 5 days. UHPLC-QTOF was used to qualitatively analyze the degradation products and intermediates (data are shown in the SI). The analysis results of the PFOS decomposition treatments detected three PFSAs, six PFCAs, five polyfluorinated acids, and six H-perfluoroalkanes. Three PFCAs (PFDA, PFOA, and PFDoA) were identified as degradation products. For PFDA, and the peak area of the degraded sample was 3.90 times the control. The peak areas of PFDoA and PFOA in the degradation samples to the controls did not double, but they were not detected in the controls. Obviously, there are both PFDA and PFDoA, whose perfluorocarbon chain length is longer than that of PFOS, and PFOA, whose perfluorocarbon chain length is shorter than that of PFOS. The formation pathway of short-chain products may involve stepwise hydrolysis after breaking the C–S bond, similar to the advanced oxidation or reduction of PFOS (Gu et al., 2017; Kim et al., 2019). It has been reported that long-chain products of PFOS can be formed in a similar biomimetic system through the stepwise addition of the C₄F₈ moiety after C–C bond cleavage (Li et al., 2020). The suspicious degradation products included PFBA and H-PFHxS. Although the peak areas of these substances significantly differed from those of the controls, they did not double. The possibility of degradation products cannot be completely ruled out. However, according to the analytical standards, H-perfluoroalkane can be identified as degradation products. It is difficult to detect H-perfluoroalkanes in the aqueous phase at the high temperatures because of their high volatility and poor solubility. The peak times of the H-perfluoroalkanes were similar to those of the corresponding PFCAs. Therefore, the H-perfluoroalkanes detected in the samples were likely to be produced by the decarboxylation of PFCAs. The remaining substances were considered as impurities contained in the standards. According to this analysis, the degradation products of PFHxS included PFHpS, PFHxA, and PFUdA, and the suspected degradation product was PFOS. The analysis of PFBS demonstrated that only PFNA was the degradation product. Because PFNA, which has the same perfluorocarbon chain length as PFOS, is difficult to degrade, to investigate the effect of the terminal functional groups, 8:2 FTOH containing –CH₂CH₂OH was introduced. The degradation products of 8:2 FTOH include PFHxA, PFOA, and PFTeDA, and the suspected degradation product is H-PFHpA.

Based on the product analysis, the proposed degradation pathways of PFOS, PFHxS, PFBS, and 8:2 FTOH are shown in Fig.4. Generally, the products types of PFOS and 8:2 FTOH were similar. Both PFOS and 8:2 FTOH, with a perfluorocarbon chain length of 8, were degraded via hydrolysis, addition of (–CF₂–)_n, C–C scission, and H/F exchange after the dissociation of terminal functional groups. However, the main degradation pathway of PFOS is that PFOS forms C₈F₁₇ radical after desulfonation, which is then hydrolyzed to form PFOA. C₈F₁₇· continuously adds C₂F₄ and then hydrolyzes to form PFDA and PFDoA. In addition, the degradation of PFOS also occurred through C–C scission and generated short-chain products, which is consistent with the results of previous studies on PFOS degradation products (Li et al., 2020; Sun et al., 2021). 8:2 FTOH mainly produced degradation products after the removal CF₂ and hydrolysis. Although the degradation products of PFOS and 8:2 FTOH were similar, the degradation mechanisms were different. The removal rate of PFOS was higher than that of 8:2 FTOH, and the defluorination rate was 25 times that of 8:2 FTOH (Fig. S1), indicating that the terminal functional groups significantly affected their removal and defluorination rates. The similarity in the types of degradation products of PFOS and 8:2 FTOH may be because they were both degraded by preferentially breaking C–C or C–S bonds. The detection of polyfluorinated acids also verified that long-chain PFAs were more completely mineralized than short-chain PFAs. The amount of degradation products increased with the growth of the perfluorocarbon chains. Moreover, the degradation products of PFHxS included PFSAs and PFCAs with lengthened carbon chains. PFCAs with lengthened and shortened carbon chains were mainly involved in PFOS. The presence of PFSAs with lengthened carbon chain in PFHxS indicated that CF₂ was added to the perfluorocarbon chain. Similarly, no short-chain products were detected during PFBS decomposition. It is possible that the addition of (–CF₂–)_n occurred before hydrolysis since short-chain PFAs have C–F bonds with high BDEs, which makes defluorination difficult. Therefore, the degradation pathways of PFBS, PFHxS, and PFOS in this study differed because of the lengths of the perfluorocarbon chain. However, the mechanism underlying the effect of PFSAs structure on the degradation pathways requires further study. Furthermore, the degradation products cultivated for only for five days were monitored in the experiment, but the type and contents of the products changed dynamically. Further research should be performed based on the relationship between the products and incubation time or extension of the incubation time.

Fig.4 Proposed degradation pathways of PFOS, PFHxS, PFBS, and 8:2FTOH. The solid arrows indicate transformation steps that are expected to occur based on degradation product analysis. The dashed arrows indicate potential transformation steps that are in doubt. The double arrows indicate multiple transformation steps that are omitted. The structural formula of polyfluorinated acids is not shown because the substitution position is uncertain.

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VB₁₂ is a transition-metal coenzyme containing cobalt at the center of its structure. Its redox state can be categorized into the trivalent stable state Co³⁺, divalent reduced state Co²⁺, and monovalent super-reduced state Co⁺. During the degradation process, the bimetallic material acts as an electron donor to reduce Co³⁺ to Co²⁺ and Co⁺. VB₁₂ becomes a strong nucleophile containing Co⁺ that attacks the C–F bond, promoting the reductive defluorination of PFAs. The intermediate is hydrogenated and dissociated into VB₁₂. This confirms that the C–F bond energy significantly affects the defluorination of PFAs.

3.4 Effect of coexisting substances

To clarify the effect of complex components in natural water on the system, PFOS, which has high defluorination and removal rate and is easier to monitor, was selected as a representative substance, and common anions and dissolved organic matter (DOM) were added to the system at different concentrations. However, the mechanisms by which PFAs are affected by coexisting substances are likely differ owing to their different degradation mechanisms. The purpose of exploring the effects of coexisting substances on PFOS is to provide a reference for future research.

3.4.1 Effect of DOM

Humic acid (HA) and fulvic acid (FA) were selected as the main components of DOM. Humus has carboxyl groups, amino groups, and carbon-oxygen double bonds, they are high reactivity (Zhang et al., 2019). Humus may react with pollutants through adsorption and coordination, thereby, affecting the removal rate in this system (Xiang et al., 2018). As shown in Fig.5(a) and 5(b), regardless of whether HA or FA was used, there was no significant impact on the removal rate of PFOS in the concentration range of 0–50.0 mg/L. The slight fluctuation in the removal rate during the reaction was attributed to the adsorption of PFOS by HA. Previous studies indicated that PFOS is highly hydrophobic and more adsorbable to HA than other PFSAs (Zhao et al., 2014; Liu et al., 2020b). A sample with an HA concentration of 100.0 mg/L was added to verify the impact of high HA concentrations. The removal rate of PFOS was significantly reduced in the presence of 100.0 mg/L of HA. In addition, the F^– concentration of the HA = 100 mg/L sample was only 182.46 μmol/L, while the sample without HA was 312.28 μmol/L, which was 1.71 times that of the sample with 100 mg/L HA added at 12 h (Fig. S2). Humus combines with metal cations precipitated in the system to produce stable complexes, which hinder the removal rate of PFOS (Dries et al., 2005). However, humus may be adsorbed by metal reductants and block surface sites (Sun et al., 2016). Additionally, they are beneficial for the degradation of PFASs under alkaline conditions. High HA concentrations make water more acidic and weaken the activity of the catalyst VB₁₂ (Sun et al., 2021). The high-concentration sample (100 mg/L) of FA showed little inhibitory effect (Fig. S3). The actual total organic carbon (TOC) concentrations produced by different humus dosages in the samples were determined (Fig. S4). The TOC contents in the samples with 100 mg/L HA and FA were 40 and 25 mg C/L, respectively. It is speculated that the effect of humus on the decomposition of PFOS may be related to the TOC content of the samples. The inhibitory effect of FA is lower than that of HA because FA is a smaller molecule; therefore, it adsorbs less reductant and occupies fewer surface sites (Dong et al., 2016). In addition, the distributions of the redox-active functional groups of HA and FA are different; FA shows a stronger reducing ability and has less impact on the reducibility of the system (Yang et al., 2016).

Fig.5 The effect of (a) HA, (b) FA, (c) SO42–, (d) Cl–, (e) DO concentration on PFOS removal; (f) DO concentration in the system after reaction. The experiment conditions: [PFOS]0 = 200 μmol/L, [VB12]0 = 200 μmol/L, [0.5 wt% nFe0/Ni0]0 = 3.0 g/L, pHi = 9.5, and t = 60 °C.

Full size|PPT slide

3.4.2 Effect of anions

Cl^– and SO₄^2– are common inorganic anions in water that may react with the reductants in this system. Therefore, the effects of various concentrations of Cl^– and SO₄^2– on PFOS removal were investigated. As shown in Fig.5(c), the removal rates of PFOS after 144 h were 45.62%, 32.03%, 30.28%, and 23.93% in the presence of 0, 0.1, 1.0, and 10 mmol/L SO₄^2–, respectively. The results showed that PFOS removal was suppressed with increasing SO₄^2– concentrations and a similar phenomenon was observed for the degradation of chlorinated organic matter by zero-valent iron materials (Liu et al., 2007). The inhibition of PFOS removal was likely caused by the accelerated dissolution of Fe⁰ and the replacement of intermediate Fe(OH)_ads by SO₄^2– to form Fe₂(SO₄)_ads (Yu et al., 2013). The complex cannot be converted quickly, prolonging the generation of Fe²⁺. It is also possible that the reaction sites on the iron surface and their oxidation products may be covered by Fe-anion complexes, affecting this effect (Song et al., 2017).

The effects of the Cl^– concentration on PFOS removal are shown in Fig.5(d). There is no obvious regular effect at low concentrations of Cl^–. PFOS removal was suppressed after adding 10 mmol/L Cl^–. However, the PFOS removal rate was 47.17%, which had only a slight impact compared to the absence of Cl^– at 144 h. Studies have indicated that a low concentration of Cl^– promotes the corrosion and shedding of the passivation layer on the surface of zero-valent metal materials, which may improve the reduction performance (Sun et al., 2016). However, when the concentration increases, competitive adsorption offsets this promotion effect and even limits the removal rates because they occupy reaction sites (Ahmad et al., 2018).

3.4.3 Effect of dissolved oxygen (DO)

This system was incubated under anaerobic conditions; however, natural water contains DO, which is a strong competitor for electrons. In addition, DO forms a dense oxide film on the surface of metallic materials, hindering contact between electrons and target pollutants (Yin et al., 2012). Therefore, it is necessary to evaluate the effect of initial DO concentration on PFOS removal. As shown in Fig.5(e), the suppression of PFOS removal by DO was concentrated during the early stages of incubation, with an initial DO of 9.5 mg/L after full aeration. The DO (Fig.5(f)) and F^– concentrations (Fig. S5) were monitored. It was observed that DO reduced the PFOS defluorination rate. The DO in the two samples reached the same level after 2 h regardless of the initial DO concentration. Combined with the fact that zero-valent metals are easily oxidized, it is suspected that DO consumes part of the bimetallic reductant first, and when the system is in an anaerobic state, the remaining reductant continues to react with VB₁₂. As the amount of reductant added was sufficient, it was only found the PFOS removal rate was affected in the first 12 h. The effects of DO were further explored under continuous oxygenation.

Most environmental factors studied had an inhibitory effect on the removal of PFAs by the system. Therefore, it is difficult to optimize the degradation effect of the system by changing the types of coexisting ions in the water. This indicates that the system may be affected when applied to water bodies with poor water quality. Only the influence of a single factor on the removal effect of the system was studied; however, an actual water body is a combination of multiple factors, which requires further exploration.

4 Conclusions

This study investigated the degradation of PFAs in a VB₁₂ + nFe⁰/Ni⁰ biomimetic reduction system and explored the influence of the self-structure of PFAs and the effect of coexisting substances in natural water. The results indicated that the defluorination and removal rates of long-chain PFAs were better than those of short-chain, and PFSAs with –SO₃– are better than PFCAs with -COOH-. The most suitable target pollutions for this system were PFCAs with C > 11 and PFSAs with C > 6. Based on products detection, it was found that the length of the perfluorocarbon chain and terminal functional groups of PFAs affected the degradation products. Thus short-chain PFAs generated only long-chain products. In particular, long-chain PFAs generate polyfluorinated acids, and their defluorination rate is high. The influence of the chemical structure of PFAs on the products is mainly reflected in the degradation pathways, and the specific mechanism needs to be further studied. Coexisting substances in natural waters primarily affect the combination of bimetallic reductant and PFAs. Both HA and FA at a concentration of 100 mg/L inhibited the defluorination and removal rates of PFOS, and the inhibition by HA was stronger at the same concentration. SO₄^2– significantly inhibits the removal of PFOS with the concentration of 0.01–10.0 mmol/L, while Cl^– showed no obvious inhibition even at 10.0 mmol/L. PFOS degradation was attenuated by DO through the consumption of bimetallic reductant. In general, the VB₁₂ + nFe⁰/Ni⁰ biomimetic reduction system is feasible for removing PFAs. However, there are also limitations, such as insufficient removal and defluorination rates and inhibition by coexisting ions. For practical applications, these conditions must be further optimized. For an anaerobic environment in the system and added bimetallic materials, the issues of maintaining anaerobic conditions and avoiding metal leakage must to be considered in actual applications. For better on-site applications, the effects of additional coexisting contaminants, optimization of system conditions, and identification of degradation pathways require further study.

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Acknowledgements

This study was supported by the National Natural Science Foundation of China (No. 51878300), the Promotion Program for Young and Middle-aged Teacher in Science and Technology Research of Huaqiao University (No. ZQN-YX602), the Science and Technology Program of Quanzhou City (No. 2018C084R) and the Experimental Teaching and Management Reform Project of Huaqiao University in 2023 (No. SY2023J03).

Conflict of Interests

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

Electronic Supplementary Material

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