Introduction

A class of fully fluorinated hydrocarbons known as perfluorocarboxylic acids (PFCAs, CnF2n + 1COOH) has been widely applied in various ranges for several decades. They are receiving increasing attention because of their easy bioaccumulation and persistent toxic environment impact. Perfluorooctanoic acid (C7F15COOH, PFOA), as a PFCA has already been detected in environment waters, human bodies and wildlife1,2,3,4. As a ubiquitous environmental contaminant, PFOA has the following features: extremely resistant degradation, bioaccumulation in food chains and long half-lives in human bodies, all of which present characteristics of persistent organic pollutants5,6. The major human exposure sources to PFOA and other PFCs include drinking water7 and edible fish8 etc and can lead to several chronic and developmental problems, such as children attention deficit/hyperactivity disorder9 and lowered immune response to vaccinations10. Statistic survey shows that its concentration in human bodies and wildlife continues to increase in certain location globally11.

However, PFOA is very stable and considered almost non-biodegradable under natural environments because of the strong C-F bonds (116 kcal/mol). Besides, some researchers reported that PFCAs could almost not be degraded by advanced oxidation process. The principal reason may be that C-F bonds can't be destroyed effectively by hydroxyl radicals (·OH)12,13. Various treatments for PFCAs including adsorption14, photocatalysis15,16, photolysis17, thermolysis18, sonochemical19 and other methods20,21 have been tested for decomposing PFCAs. However, harsh reaction conditions at high temperatures and high pressures are usually needed18,19. Furthermore, the mineralization and defluorination of PFCAs always could not be achieved completely and toxic by-products might be formed during the decomposing processes12. Thus, it is highly desirable not only to decompose the PFCAs, but also to defluorinate the toxic by-products for complete mineralization of PFCAs. A novel method for PFCAs mineralization with high efficiency is still highly desired.

It is well-known that γ-ray irradiation can also degrade persistent organic pollutants efficiently owning to the production of active species, e.g., ·OH and eaq−, which make γ-ray irradiation a promising method for wastewater and sludge treatment22,23. Also, it has several unique advantages and especially it needs no additional chemicals to fully complete mineralization of pollutants24,25,26. When aqueous solutions are irradiated with γ-ray, water will absorb most of the radiation and produce several activated species, in which ·OH and eaq− are predominant, as shown in Eq. 1. The numbers in brackets are the yields (μmol/J) of those species per unit of radiation at pH 7.027. However, it has not been reported whether PFCs could be mineralized efficiently by γ-ray irradiation or not.

Therefore, in this work, the efficiency of the radiation-induced degradation and defluorination of PFOA in aqueous solution by a 60Co γ-source was explored. The mineralization mechanisms of PFOA and the crucial roles of ·OH and eaq− generated were also investigated. To be the best of our knowledge, this might be the first report about the degradation of PFOA by γ-ray irradiation. The results would bring an efficient and environmentally friendly technique for PFCAs degradation.

Results and Discussion

Mineralization of PFOA by γ-ray irradiation

The degradation of PFOA by irradiation at various pH values is shown in Figure 1. The degradation efficiency of PFOA increased with the increasing pH value. At pH < 7.0, PFOA could not be decomposed efficiently. However, at pH 13.0 the degradation efficiency could reach almost 100% after 6-h irradiation. Also, the defluorination efficiency was also nearly 100% at pH 13.0 (Figure 1B). Furthermore, the released F− concentration also decreased with the decreasing pH. At pH < 7.0, almost no F− could be detected. This might be attributed to the loss of eaq− through Eq. 2:

Figure 1
figure 1

(A) PFOA degradation by γ-ray at various pHs. The dotted line shows the change of TOC of the solution during irradiation at pH 13.0; and (B) Concentrations of F− released into the solution at various pHs.

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eaq− (E0 = −2.9 eV) was a much stronger reducing agent than H· (E0 = −2.1 eV) and might play a more important role in the degradation of PFOA rather than H·.

Total organic carbon (TOC), as an important index of mineralization, decreased from 4.2 mg/L to 0.6 mg/L after 22-h irradiation at pH 13.0 (Figure 1A). The mineralization degree of PFOA was around 86%, meanwhile the defluorination ratio was nearly 100%, indicating that most of the PFOA was degraded to CO2, fluorine ion and water. The high efficiency of defluorination and TOC removal implies that nearly all the C-F and C-C bonds in PFOA could be broken by γ-ray irradiation.

High degradation kinetics of PFOA by γ-ray irradiation

To compare the degradation rate of PFOA by γ-ray irradiation at various conditions, a pseudo-first-order kinetic model was used to describe the degradation kinetics:

where C and C0 are the representative PFOA concentrations at time t and at initial. k is the apparent first-order rate constant. The linear regression between ln(C/C0) vs t under various conditions is shown in Figure 2. The high correlation coefficiencies at various pHs indicate that the pseudo-first-order kinetic model was appropriate to describe the degradation of PFOA by γ-ray irradiation.

Figure 2
figure 2

PFOA degradation kinetics at different pHs.

The lines were regressed using the pseudo-first-order kinetic model.

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The degradation of PFOA could be promoted in an alkaline medium. As Table 1 shows, the linear fit of the kinetic plots gives k7 = 0.034 h−1, k9 = 0.043 h−1, k11 = 0.199 h−1 and k13 = 0.654 h−1 at pH 7.0, 9.0, 11.0 and 13.0, respectively. The high PFOA degradation rate constant at pH 13.0 was about 20 times of that at pH 7.0 and 15 times of that at pH 9.0. Furthermore, the degradation rate constants in this study were much higher than those reported previously by other methods28,29,30, indicating that the γ-ray irradiation is efficient for the PFOA decomposition.

Table 1 The pseudo-first-order apparent degradation rate constants and half times of PFOA by γ-ray irradiation at various pH values
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Roles of eaq− and ·OH in the PFOA degradation

Preliminary experimental results show that PFOA is degraded and defluorinated efficiently in alkaline solution under N2-saturated conditions. The most likely active species responsible for the degradation of PFOA are ·OH and eaq− according to Eq. (1). To investigate the contributions of the activated species produced by the γ-ray on the degradation of PFOA, PFOA samples were irradiated with the dose of radical scavengers (e.g., t-BuOH and H2O2), under N2- or O2-saturated conditions respectively (Table 2). The dose of scavenger was able to confirm which activated species was crucial in the PFOA decomposition or defluorination. As shown in Table 2, after 22-h reaction, in the presence of H2O2 or O2, PFOA seemed to be resistant to γ-ray irradiation and no degradation or defluorination of PFOA could be detected after irradiation. This implies that the radicals eliminated by H2O2 or O2 were essential for the degradation and defluorination of PFOA. It is well known that the reducing radicals eaq−, would react with O2 and H2O2 through Eqs. 5 and 6 and turned into weak activated radicals O2−· and HO2·. Consequently, eaq− should be the crucial radical for the cleavage of both C-C and C-F bonds and PFOA can't be degraded by ·OH radical only.

Table 2 Degradation efficiency of PFOA after 22-h irradiation at different conditions
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t-BuOH, a ·OH scavenger, was added to the PFOA solutions before irradiation. As shown in Figure 3, with the increasing t-BuOH concentration, less F− was released. With the addition of 100 mmol/L t-BuOH, little F− could be detected after irradiation. However, it was surprising to find that the degradation efficiency of PFOA was nearly 100% no matter how much t-BuOH was dosed. These results imply that C-F bond could not be broken efficiently in the absence of ·OH. With the decrease in t-BuOH dosage, the amount of ·OH in the aqueous sample gradually increased and more F− could be found in the solution after irradiation. These results show that both ·OH and eaq− were responsible for the degradation of PFOA by γ-ray irradiation.

Figure 3
figure 3

Effects of addition of t-BuOH on the degradation and defluorine of PFOA.

The samples were irradiated for 12 hours in N2-saturated atmosphere and at pH 13.0.

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Mechanism of PFOA decomposition by irradiation

The decomposition products of PFOA in the aqueous solutions were identified using LC-MS method and the results are shown in Figure 4, while the mass spectra of various intermediated products are shown in Figure S1. It was observed that the concentrations of PFHpA (C6F13COOH), PFHxA (C5F11COOH) and other intermediate products increased at initial after irradiation, but decreased gradually. After 22-h degradation, all these intermediate products were degraded.

Figure 4
figure 4

LC-MS of the intermediated products of shorter-chain perfluorocarboxyl anions including C3F7COO− (PFBA), C4F9COO− (PFPeA), C5F11COO− (PFHxA) and C6F13COO− (PFHpA) at various irradiation times.

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On the basis of the products identified, the possible degradation pathway of PFOA was proposed and is illustrated in Figure 5. DFT simulations were used to investigate the thermodynamic performance for the reactions of eaq− and ·OH at different sites on PFOA molecule. Configurations of the initial reactants and intermediate products are shown in Figure S2. And the structural parameters of the reactive sites through the degradation process are presented in Table S1. PFOA (C7F15COO−) was considered to be degraded to short-chain PFHpA (C6F13COO−), through a series of reactions (Table 3). At the beginning, the eaq−-mediated reduction occurs at the carbonyl group with a subsequent loss of HO− and CO, yielding an alkyl radical, or in this case a fluoroalkyl radical (C6F13CF2·, Table 3). The negative Gibbs free energy change (ΔGθ = ΔG1θ + ΔG2θ + ΔG3θ = −3.059 eV) of the initial three steps indicates that the generation of C6F13CF2· from C7F15COO− is thermodynamically favored according to the DFT calculations (Table 3). The bond lengths of C7-F14 and C7-F15 decrease with the bond angle of F14-C7-F15 increasing (Table S1). Following this, the radical C6F13CF2· in aqueous solution will react with ·OH to form unstable alcohol C6F13CF2OH, which undergoes HF elimination to form C6F13COF with ΔG5θ of −0.116 eV. After that, ·OH will attack and cleave the C = O bond to form another radical C6F13CF(OH)O·, which can be easily attacked by eaq−. Then, PFHpA (C6F13COO−) is formed by releasing the HF. The generated PFHpA can also undergo this process to be degraded to PFHxA (C5F11COO−), as shown in Figure 5. By repeating this CF2-unzipping cycle, PFOA could be completely decomposed to CO2 and F−.

Table 3 Calculated thermodynamic characteristics of the PFOA degradation by radical eaq− and ·OH in aqueous solution generated through γ-ray irradiation
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Figure 5
figure 5

Proposed PFOA degradation pathway with synergies of ·OH (O·−) and eaq− generated by γ-ray irradiation.

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In the alkaline aqueous solution, ·OH may be converted to O·− for the pKa value of ·OH is 11.931 and H+ was scavenged by OH−. The thermodynamic properties for the reactions of eaq− and O·− with PFOA molecules were calculated and listed in Table S2. The steps 2, 4–7 of ·OH degradation pathway in Table 3 were replaced by the steps O2 and O4 ~ O7 in Table S2 in the Supplementary information. The DFT calculation indicates that the PFOA degradation could also occur spontaneously in the alkaline solution with the synergistic effects of eaq− and O·− generated by γ-ray irradiation. The possible degradation pathway of PFOA by O·− and eaq− is also illustrated in Figure 5.

Furthermore, some possible alternative reactions for PFOA degradation are also investigated, as listed in Table S3. Configurations of the possible intermediate products are shown in Figure S3. The thermodynamic properties of PFOA degradation individually by radical eaq− or ·OH in aqueous solution were calculated. The results agreed well with the experimental results, i.e., only either eaq− or ·OH might not be able to accomplish complete mineralization of PFOA.

Limitations and future research

Use of γ-ray irradiation for wastewater and contaminated sludge treatments is a promising treatment technology22,23. The disinfection of sludge for beneficial reuse has been demonstrated at full scale using γ-ray irradiation32. And a lab scale wastewater treatment system using γ-ray irradiation directly is being developed33,34,35. But people are sensitive about the safety of γ-ray irradiation used for water treatment. The radiation leaks may affect health issues for γ-rays are biologically hazardous. Therefore, protection and safety measures for sources must be made in a way that they can be regularly monitored and verified for safety and the consummate management ensures the safety of radiation application for wastewater treatment. There are a number of features that make γ-ray irradiation an attractive process to pursue for eventual application. For example, the process requires no additional chemicals and does not leave any hazardous residues (chemical or radioactive). Both oxidizing and reducing reactive species, such as eaq− and ·OH, are formed in relatively the same proportion in aqueous media. Moreover, the process is rapid and logistically simple, allowing for maximum flexibility in plant design. In future researches, as a promising approach, γ-ray irradiation will be applied to facilitate the degradation of other persistent organic pollutants, such as polychlorinated biphenyls etc. Furthermore, radiation processing using electron beam accelerator or alpha radiation with the same advantages as γ-ray irradiation, but controllable and safe, could be used in the wastewater treatment for achieving practical industrialization.

Conclusion

In summary, the complete degradation of PFOA could be achieved by γ-ray irradiation in this work. Results show that the two radicals, ·OH and eaq−, played the most important roles in the complete mineralization of PFOA. The proposed mechanism of PFOA degradation by eaq− and ·OH were elucidated by both experimental measurements and theoretical calculations. The synergistic roles of ·OH and eaq− involved in the cleavage of C-C and C-F bonds and therefore complete mineralization of PFOA could be achieved. This work provides a promising approach by virtue of its environmentally friendly and complete mineralization of PFOA, which will facilitate the practical application of PFCAs by γ-ray irradiation in wastewater treatment.

Methods

Reagents

PFOA (>98%) was purchased from Aladdin Corp. H2O2, tertiary butanol (t-BuOH) and other chemicals were obtained from Shanghai Chemical Reagent Co., China, which were all analytical grade (>99.7%). All aqueous solutions were prepared with high purity water produced by a Millipore water purification system (Millipore Inc., USA).

Irradiation experiments

PFOA solutions with an initial concentration of 20 mg/L were kept in 100 mL Pyrex glass vessels sealed with teflon caps and then were irradiated by 60Co-γ-source at an irradiation dose rate of 96 Gy/min. In preliminary experiments, PFOA solutions with three different pHs were prepared and irradiated. The effects of the atmosphere of solutions were also evaluated by saturated with either N2 or O2 gas. To identify the contributions of the activated species to PFOA degradation, the radical scavengers (t-BuOH, or H2O2) were added to PFOA solutions prior to irradiation. The experimental designs are listed in Table 2. All experiments were conducted at ambient temperature in parallel.

PFOA measurement and other analyses

Concentrations of PFOA and its degradation products were determined by a liquid chromatograph (Accela 600Bar, Thermo Co., USA) coupled with a mass spectrometric system (LTQ-Orbitrap, Thermo Co., USA) (LC-MS). Prior to analysis, each sample was diluted 15 times with methanol to ensure the PFOA concentration below 2 mg/L. The external method was used for the quantitative analysis of PFOA. Solutions with various PFOA concentrations were prepared to make the external calibration curve in the range of 0–2 mg/L. The separation column was a Hypersil GOLD C18 column (2.1 mm i.d. × 100 mm, 3 μm particles) and the column temperature was set at 40°C. The flow rate was maintained at 0.2 mL/min with a mobile phase of eluent A (10 mM ammonium acetate in water) and B (methanol). The eluent gradient started with 40% B for 1 min and then was linearly increased to 100% B in 6 min, kept for 5 min and finally returned to the starting conditions within 1 min and held for 2 min for equilibration during injection interval. MS detection was operated in negative mode by using an electrospray ionization source. The parameters were optimized for maximal transmission of the PFOA anion (m/z 413) resulting in the following instrument parameters: nitrogen sheath gas flow rate, 20arbitrary units; Aux/sweep gas flow rate, 5 arbitrary units; spray voltage, 4 kV; heated capillary temperature, 275°C.

The concentration of fluorine ion (F−) in aqueous samples was measured by an ion chromatograph system (ICS-1000, Dionex Co., USA), which was equipped with a pump, a degasser, a guard column and a separation column (Dionex IonPac AS 14A, 4 mm i.d. × 200 mm). The sample injection volume is 25 μL and the mobile phase was NaHCO3 (1 mmol/L) and Na2CO3 (8 mmol/L). The flow rate was set at 1 mL/min. The suppressor current was set at 124 mA.

TOC was used to evaluate the mineralization degree of PFOA during γ-ray irradiation, which was measured by using a TOC analyzer (VCPN, Shimadzu Co., Japan).

Quantum chemical calculations

Quantum chemistry method can offer an effective and rapid alternative to elucidate the reaction mechanism of refractory organic pollutants36,37,38,39. The possible PFOA mineralization mechanisms by γ-ray irradiation in aqueous solution were studied using spin-unrestricted DFT computations. In the calculation, an all-electron method within the Perdew, Burke and Ernzerhof (PBE) form of generalized gradient approximation (GGA)40 for the exchange-correlation term and the double precision numerical basis sets including p polarization (DNP)41 were employed, as implemented in the DMol3 code42. The energy in each geometry optimization cycle was converged to within 1 × 10−5 Hartree with a maximum displacement and force of 5 × 10−3 Å and 2 × 10−3 Hartree/Å, respectively. The solvent effect of the water medium was described using the conductor-like screening model (COSMO)43. COSMO is a continuum solvent model where the solute molecule forms a cavity within the dielectric continuum of permittivity that represents the solvent.