Introduction

Per- and polyfluoroalkyl substances (PFASs) are a large class of synthetic fluorinated chemicals that can be always found in nonstick cookware, fire retardants and waterproof clothes, with perfluorooctanoic acid (PFOA) as a representative substance1. It raised public concerns recently due to their ubiquitous presence, environmental persistence, potential carcinogenicity and toxicity2,3,4,5,6. According to recent human epidemiology studies, long-term exposure to PFASs leads to numerous noncancer health effects, including increased liver enzymes and cholesterol7,8, decreased birth weight and fetal growth9,10,11, as well as decreased vaccine response8,12. Increasing kidney and testicular cancer were proved to be associated with exposure to PFOA in occupationally exposed workers and communities near PFASs contamination sources13,14,15. In 2024, the United States Environmental Protection Agency (USEPA) has posted the maximum contaminant levels (MCLs) of 4 ppt individually for PFOA and perfluorooctanesulfonic acid (PFOS), along with health-based goals of 0 ppt for these two PFASs, and 1.0 Hazard Index for the combined concentrations of perfluorohexanesulfonic acid (PFHxS), hexafluoropropylene oxide dimer acid (HFPO-DA), perfluorononanoic acid (PFNA) and perfluorobutanesulfonic acid (PFBS)16.

The unique structure of PFASs that part or all hydrogens on carbon are replaced by fluorine atoms makes it hard to be degraded by conventional wastewater treatment technologies17. A few innovative techniques have been developed to fully mineralize PFASs, such as sonochemical degradation18,19, plasma-based technology20,21, alkaline hydrothermal treatment22,23, advanced reduction processes24,25 and biodegradation26,27. Electrochemical treatment is a main research wastewater treatment technology that attracts much interest because of their capacity of completely mineralizing persistent organic pollutants (POPs)28,29. The advantages of electrochemical treatment include high energy-efficiency, facile scalability, and can be conveniently set up and operated at mild conditions30,31. Electrooxidation (EO) can destruct POPs via direct electron transfer (DET) on anode surface and reactive radical species mediated reaction under room temperature and pressure32,33. Recent studies have shown that boron-doped diamond (BDD) anode had the capability to remove PFASs due to its high oxygen evolution potential (OEP) and reactivity30,34,35, 99.5% ± 0.1% PFOA removal and 17.7% ± 0.7% defluorination were achieved after 4-hr EO treatment on BDD anode at 21.4 mA·cm−2 when initial concentration of PFOA was 1 mg·L−134.

Photolysis has also emerged as one of the most promising technologies in PFASs treatment due to its chemical-free processes and synergistic capacity with other treatment methods36,37. Vacuum ultraviolet (VUV) irradiation is a highly efficient method for water treatment due to in situ generation of hydroxyl radicals38. It can effectively remove pesticides39, antibiotics40, resistance genes41 and micropollutants42. As reported, the removal rate of pollutants by VUV was 1.3–57 times that of ultraviolet (UV)43. However, due to strong C-F bond, direct photolysis of PFASs was not very effective, 61.7% removal of PFOA with 17.1% defluorination ratio was achieved after 2-hr treatment under 185-nm VUV irradiation, while only 33% removal of PFOA was reported for treatment at 254-nm UV when the initial concentration of PFOA was 41.4 mg∙L−136. In order to improve the degradation efficiency of photolysis on PFAS compound without addition of extra chemicals, we hypothesize that PFASs degradation can be accelerated via combining EO with photolysis processes.

This study aims to investigate the effect of VUV irradiation on EO treatment of model PFAS compounds, PFOA and perfluorohexanoic acid (PFHxA). BDD electrode was selected as the anode material. Quenching experiments were employed in combination with Density Functional Theory (DFT) computation to elucidate the mechanisms of the PFASs degradation in EO system in the presence of VUV irradiation. This study also examined the feasibility of photo/electrochemical method on removing PFASs from real industrial wastewater.

Materials and methods

A detailed description of chemicals and reagents used in this study is provided in the Supporting Information Text S1.

Physicochemical and electrochemical characterizations

The crystalline phases of the BDD anode were identified using an ARL EQUINOX 3000 X-ray diffractometer (XRD) (Thermo Fisher, USA) with CuKα1 radiation at 45 kV/40 mA. Surface morphology was scanned using a Hitachi Regulus-8100 Scanning electron microscopy (SEM) (Tokyo, Japan). Anodic potential (AP) measurement, linear sweep voltammetry (LSV) and limiting current technique were performed using a CHI-660E electrochemical workstation (Shanghai, China) with a leak-free Ag/AgCl reference electrode (CH Instruments, Inc., CHI111). All potentials were adjusted with internal resistance (IRs) compensation and reported versus the standard hydrogen electrode (SHE).

Industrial wastewater sample collection and storage

The industrial wastewater samples were collected from the water inlet of industrial wastewater treatment plant of a chemical industry park in Jiangsu Province, China. Samples were transported to Jiangsu Provincial Key Laboratory of Environmental Engineering in 1-L high-density polyethylene (HDPE) containers containing ice packs. The samples were stored at 4 °C until characterization and experimentation.

VUV and electrochemical experiments

Batch experiments using spiked reaction solution or industrial wastewater were conducted using a 500-mL polypropylene beaker containing 400-mL reaction solution. The BDD plate (coated on both sides, 10 cm ✕ 5 cm, New Frontier Technology Co., Ltd., China) and two titanium plates of the same size and shape were placed at 0.50-cm gap as the anode and cathodes, respectively. Constant current density was applied to the reactor using a controllable DC power source (Kangkesi Inc, China). The immersion area of anode was 20 cm2 (single side). VUV radiation source was a 20 W 185-nm low-pressure Hg lamp, which was obtained from Xingchuang Scientific Inc. (Shenzhen, China). The schematic of the reactor and the emission wavelength of lamp were shown in Figure S1A and Figure S1B, respectively. The main emission wavelength of the Hg lamp is 185 nm VUV. The whole reactor was placed in a water bath to keep the temperature stable during the whole experiments (25.0 ± 3.0 ℃). The EO treatment with and without VUV irradiation were named EO and EO + VUV in this study, respectively. The reaction solution contained a mixture of PFOA and PFHxA (1.00 mg·L−1 each), and 100-mM Na2SO4 as supporting electrolytes.

Chemical analysis

For experiments on spiked reaction solutions, duplicate 400-µL samples were taken at each prescribed interval for chemical analysis. A 400-µL aliquot was mixed with 400-µL methanol containing 80-ppb M8PFOA and M5PFHxA (for spiked solution), filtered through a 0.22-µm nylon filter. Duplicate 1 mL samples were collected at predetermined reaction time for industrial wastewater experiments. 10 µL internal standard mixture (EPA-533ES from Wellington Laboratories, Canada) was added to each sample. Wastewater samples were filtered through a 0.22-µm polypropylene (PP) membrane to remove suspended solids. All samples were stored at 4℃ for subsequent analysis of PFASs.

PFAS quantification was performed on an ultra-performance liquid chromatography coupled with a triple-stage quadrupole mass spectrometer (Thermo Fisher Ultimate 3000-TSQ Fortis Plus), with the UPLC method summarized in Table S1 and described in Text S2 with QA/QC protocols. The transition and limit of detection of all target PFASs analyzed in this study are listed in Table S2. The target PFASs and their correspondent isotope-labeled internal standards are summarized in Table S3. Total organic carbon (TOC) concentrations were measured by a TOC-L CPH Total Carbon Analyzer (Shimadzu Corp., Japan) using a catalytic combustion method at 600 °C.

The fluoride ion analysis was performed using an ion chromatography (IC) system (882 Compact IC plus, Metrohm, Switzerland) equipped with a high-capacity selective ion chromatography analytical column (Metrosep A Supp 5–150/4.0, Metrohm, Switzerland). The defluorination ratio (Fr) was calculated by dividing the released F− concentration with the total fluorine in the PFOA and PFHxA that have been removed from the system. More information about F− quantification and the calculation of Fr is provided in Text S3.

Quantum mechanics simulations

The DFT calculations were performed using Gaussian software at the PBE0/def2-TZVP level of theory, incorporating the implicit solvent model for water. After the structural optimization, a frequency analysis was conducted to confirm convergence to a stationary point on the potential energy surface with no imaginary frequencies, indicating a minimum energy structure. Time-Dependent Density Functional Theory (TDDFT) was employed for the excited-state calculations, taking 20 states into account for each structure.

Results and discussion

Characterization of BDD anode

XRD was performed to characterize BDD anode used in this study, the diffraction peaks at 43.92° and 73.29° correspond to lattice plane (111) and (220) of diamond (Figure S2A). The SEM result shows that the grain size of BDD anode is around 5–20 μm (Figure S2B).

LSV result (Fig. 1) reports the LSV scans of the BDD anode in 100 mM Na2SO4 solution with/without the presence of VUV irradiation. H2O molecules can strongly absorb light in the VUV range to form hydroxyl free radicals (\(\:HO\bullet\:\))38. VUV may provide extra energy for water oxidation so that the oxygen evolution reaction can proceed at lower anodic potential (~ 2.27 V vs. SHE) compared to pure EO system (~ 2.32 V vs. SHE). The oxygen evolution potential of the BDD anode used in this work was consistent with that in earlier studies (around 2.3 V vs. SHE)44,45.

Fig. 1
Fig. 1
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LSV scan of BDD anode.

Photo- and electro-chemical degradation of PFOA and PFHxA

Figure 2A displays the degradation curve of PFOA and PFHxA during EO experiments at 10 mA∙cm−2 with and without VUV irradiation (EO and EO + VUV). A photolysis treatment was carried out without electricity to examine the degradation ability of VUV irradiation on perfluoroalkyl carboxylic acids (PFCAs). When the reaction solution was exposed to VUV irradiation, no appreciable removal of PFOA (P-value = 0.973) or PFHxA (P-value = 0.859) were observed within 8 h (Fig. 2A), as the result of ANOVA tests (Text S4) indicated. The results were contradictory to previous findings that VUV irradiation itself can lead to PFOA decomposition without the presence of a photocatalyst36,46. However, the initial concentration of PFOA used in this study (2.42 µM) was much lower than that in previous studies (0.1 mM36 and 1.35 mM46). VUV light was strongly absorbed by water molecules and the 90% radiation power was faded over a thickness of 5.5 mm in the water38, thus photolysis of PFOA was not evident within 8 h in this work. In either EO or EO + VUV system, PFOA degraded faster than PFHxA, consistent with our previous study47. PFOA/PFHxA degradation increased in the presence of VUV irradiation during EO processes. After 2-hr EO treatment, 39.7% PFHxA and 79.2% PFOA were removed from the reaction solution at 10 mA∙cm−2, while in the presence of VUV irradiation, 54.6% PFHxA and 88.4% PFOA were degraded at the same current density, respectively. As shown in Fig. 2A, the degradation curves of PFOA and PFHxA appeared to follow pseudo-first order kinetic model (Text S5), and the observed reaction rate constant \(\:{\text{k}}_{\text{obs,PFAS}}\) was calculated via data fitting (Fig. 2B). It can be further normalized to anode surface area to obtain surface area normalized reaction rate constant \(\:{\text{k}}_{\text{SA,PFAS}}\) (Text S5). The \(\:{\text{k}}_{\text{obs,PFAS}}\) of PFOA and PFHxA increased from \(2.31\times10^{-4}\pm2.77\times10^{-6}\) s−1 and \(9.17\times10^{-5}\pm5.50\times10^{-7}\) s−1 to \(3.17\times10^{-4}\pm2.67\times10^{-6}\) s−1 and \(1.56\times10^{-4}\pm2.14\times10^{-6}\) s−1 when VUV was introduced into the system, respectively.

Fig. 2
Fig. 2
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A The concentration profile of PFOA/PFHxA versus time of a treatment with either VUV irradiation or 10 mA∙cm−2 current density (EO), or both VUV and 10 mA∙cm−2 current density (EO + VUV). B Plot of kinetic fit curve of PFOA/PFHxA degradation in EO (10 mA∙cm−2) and EO + VUV (10 mA∙cm−2) system. Initial PFOA/PFHxA concentration: 1.00 mg/L, supporting electrolyte: 100-mM Na2SO4. Error bar represents standard deviations of replicates.

Electrochemical oxidation coupled with VUV irradiation of PFOA and PFHxA was evaluated at five different current densities ranging from 5 mA·cm−2 to 25 mA·cm−2 in the batch reactor, EO experiments at same current densities were also conducted for comparison (Figure S3). The anodic potential corresponding to each current density was measured and listed in Table S4 and Table S5, which ranges from 3.27 to 4.82 V vs. SHE. Anodic potential is a vital parameter in EO treatment, since it is the driving force of direct electron transfer of target compounds and OH• generation48,49. The \(\:{\text{k}}_{\text{obs,PFAS}}\) and \(\:{\text{k}}_{\text{SA,PFAS}}\) values of PFOA and PFHxA in either EO or EO + VUV system are summarized in Table S4 and Table S5, respectively, and the \(\:{\text{k}}_{\text{SA,PFAS}}\) versus anodic potential was displayed in Fig. 3A. Limiting current technique was used to measure the mass transfer rate of PFOA and PFHxA on the BDD anode (Text S6, Figure S4)50. The mass transfer rate of PFOA and PFHxA were \(1.17\times10^{-4}\) m·s−1 and \(1.42\times10^{-4}\) m·s−1, indicating that all the reactions studied in this work were kinetically controlled. As shown in Fig. 3A, the surface area normalized reaction rate constants \(\:{k}_{SA,PFAS}\) in EO + VUV system were much higher than those in EO system when the anodic potential was between 3.27 and 4.82 V vs. SHE. The \(\:{k}_{SA,PFOA}\) of PFOA were \(1.05\times10^{-5}\pm2.95\times10^{-8}\) m∙s−1, \(2.31\times10^{-5}\pm2.77\times10^{-7}\) m∙s−1, \(2.51\times10^{-5}\pm5.89\times10^{-8}\) m∙s−1, \(3.24\times10^{-5}\pm2.38\times10^{-7}\) m∙s−1 and \(3.58\times10^{-5}\pm6.64\times10^{-7}\) m∙s−1 at 5 mA∙cm−2, 10 mA∙cm−2, 15 mA∙cm−2, 20 mA∙cm−2 and 25 mA∙cm−2 in the absence of VUV irradiation, respectively, raised to \(2.77\times10^{-5}\pm6.68\times10^{-7}\) m∙s−1, \(3.17\times10^{-5}\pm2.67\times10^{-7}\) m∙s−1, \(3.42\times10^{-5}\pm1.01\times10^{-6}\) m∙s−1, \(4.48\times10^{-5}\pm2.34\times10^{-7}\) m∙s−1 and \(5.77\times10^{-5}\pm2.18\times10^{-6}\) m∙s−1 at 5 mA∙cm−2, 10 mA∙cm−2, 15 mA∙cm−2, 20 mA∙cm−2 and 25 mA∙cm−2 when VUV was added to the EO system, respectively. PFHxA exhibited a similar trend, the \(\:{k}_{SA,PFHxA}\) increased 183%, 69.7%, 91.2%, 60.3% and 53.4% at 5 mA∙cm−2, 10 mA∙cm−2, 15 mA∙cm−2, 20 mA∙cm−2 and 25 mA∙cm−2 in EO+VUV system compared to that in EO only system, respectively. The concentration of fluoride ion was measured at the end of each treatment. The theoretical and measured values of F− concentration were listed in Table S6. When the current density was 10 mA∙cm−2, the defluorination ratio increased from 28.5 ± 2.47% to 48.6 ± 2.15% when VUV was introduced to the system. Compared to the defluorination ratio in EO system, EO+VUV exhibited greater defluorination rate at the same current density (Fig. 3B and Table S6). It suggests that VUV irradiation can greatly enhance the mineralization of PFOA and PFHxA in EO system.

Fig. 3
Fig. 3
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Surface area normalized rate constant (A) and defluorination ratio (B) in relation to the anodic potential (vs. SHE) and current density for PFOA and PFHxA degradation on BDD anode in EO and EO + VUV system, respectively. Initial PFOA/PFHxA concentration: 1.00 mg/L, supporting electrolyte: 100-mM Na2SO4. Error bar represents standard deviations of replicates. The reaction time of experiments with current density 5 mA∙cm−2, 10 mA∙cm−2, 15 mA∙cm−2, 20 mA∙cm−2 and 25 mA∙cm−2 were 9 h, 8 h, 8 h, 6 h and 6 h, respectively.

The effect of the initial pH on PFOA/PFHxA degradation in EO + VUV system was evaluated at pH 3.24–10.17 and the results are displayed in Figure S5. Acidic conditions slightly favored PFOA/PFHxA degradation. The PFOA removal rate of 83.1%, 80.6% and 78.2% were achieved after 1-hr EO + VUV treatment at the pH of 3.24, 6.20 and 10.17. PFHxA exhibited similar trends. It might be due to excess HO− in alkaline solution occupying the reactive sites on anode surface, resulting in the hinderance of PFOA/PFHxA degradation44,51,52.

Degradation mechanism of PFOA/PFHxA in EO + VUV system

 As shown in Figure S6, it is well-accepted that the degradation of PFCAs is initiated by direct electron transfer (DET) process via (R-1)53, followed by indirect oxidation pathway where PFCA radicals were attacked by highly reactive species generated on anode surface, such as hydroxyl free radicals (\(\:HO\bullet\:\)) via (R-2), (R-3), (R-4) and (R-5) to form shorter-chain PFCAs35,54,55. Then shorter-chain PFCAs continue to undergo decarboxylation and \(\:-{\text{CF}}_{\text{2}}-\) unzipping cycle, and finally mineralize to CO2 and F−49,56.

$$\:{C}_{n}{F}_{2n+1}{COO}^{-}\to\:{C}_{n}{F}_{2n+1}COO\bullet\:+{e}^{-}$$
(R-1)
$$\:{C}_{n}{F}_{2n+1}COO\bullet\:\to\:{C}_{n}{F}_{2n+1}\bullet\:+{CO}_{2}$$
(R-2)
$$\:{C}_{n}{F}_{2n+1}\bullet\:+HO\bullet\:\to\:{C}_{n}{F}_{2n+1}OH$$
(R-3)
$$\:{C}_{n}{F}_{2n+1}OH\to\:{C}_{n-1}{F}_{2n-1}COF+{H}^{+}+{F}^{-}$$
(R-4)
$$\:{C}_{n-1}{F}_{2n-1}COF+{H}_{2}O\to\:{C}_{n-1}{F}_{2n-1}{COO}^{-}+{2H}^{+}+{F}^{-}$$
(R-5)

The molecular structures and electrostatic potential (ESP) images of PFOA and PFHxA were shown in Figure S7. TD-DFT was employed to compute the UV–vis spectrums of PFOA and PFHxA ions, and the results were displayed in Fig. 4. PFOA and PFHxA ions can only absorb light below 250 nm, the highest molar extinction coefficient (ε) of PFOA and PFHxA ions fell into the range of 135 nm to 160 nm, which was within the emission wavelength range of the Hg lamp used in this work, suggesting that PFOA and PFHxA ions can absorb the VUV light for further degradation reactions. As shown in Tables 1 and 2, excited states of PFOA and PFHxA with excitation wavelength greater than 135 nm, corresponding to the shortest emission wavelength of the lamp used in this study, were listed. The first excited state of PFOA and PFHxA at 216.25 nm and 217.60 nm are mainly due to electron density moving from HOMO-1 and HOMO to LUMO, respectively. The orbital phase of PFOA and PFHxA was shown in Figure S8 and Figure S9, respectively. The ground state PFOA and PFHxA ions absorbed the VUV irradiation to generate excited state form. Excited state of molecules exhibited higher energy and reactivity than the ground state form57,58. VUV-induced excitation facilities the electron loss from PFOA and PFHxA ions to the BDD anode by reducing energy required for direct electron transfer56, which suggests that the excited state PFOA and PFHxA ions are more likely to lose an electron to generate PFOA and PFHxA radicals than that of the ground PFOA and PFHxA ions56. Direct electron transfer reaction is considered to be the rate-limiting step in PFCAs degradation30,47,49, thus VUV irradiation can promote PFOA and PFHxA degradation in EO system mainly due to accelerated direct electron transfer process.

Fig. 4
Fig. 4
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The molar extinction coefficient of PFOA and PFHxA ions obtained from Gaussian 16 calculation.

Table 1 Predicted excited States of PFHxA molecule by TDDFT calculation.
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Table 2 Predicted excited States of PFOA molecule by TDDFT calculation.
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Quenching experiment was also conducted to examine the effect of \(\:HO\bullet\:\) on PFOA/PFHxA degradation using methanol as the \(\:HO\bullet\:\) scavenger59. 0.2 M and 0.02 M methanol were added to the EO + VUV system respectively and the results were presented in Figure S10. It is interesting that the methanol concentration had no effect on PFOA degradation, however, the degradation of PFHxA was delayed when methanol was added. After (R-1) and (R-2) reactions (without \(\:HO\bullet\:\) involvement), PFOA cannot be detected in the EO + UV system. It suggests that PFOA degradation rate was mainly controlled by the rate of direct electron transfer process (R-1) instead of \(\:HO\bullet\:\) concentration in the reaction solution, which was in agreement with previous studies30,47,49,56. However, PFHxA is an intermediate product of PFOA degradation, PFOA can be degrade to PFHxA by losing two electron and consuming two hydroxyl free radicals (\(\:HO\bullet\:\)) via (R-1), (R-2), (R-3), (R-4) and (R-5), so that PFHxA concentration may increase during EO + UV system in the presence of PFOA. Reduced \(\:HO\bullet\:\) concentration, caused by methanol addition in EO + UV system, may decelerate the mineralization process of both PFOA and PFHxA, hindered the (R-3) reaction, which can be only reflected on the PFHxA degradation. It indicates that hydroxyl free radicals (\(\:HO\bullet\:\)) play a certain role in PFCAs degradation in EO + UV system. It has been proved that water can absorb VUV radiation and thus generate \(\:HO\bullet\:\) via (R-6)38, so that extra \(\:HO\bullet\:\) formation in EO + UV system may promote PFOA and PFHxA degradation, it can also further verified by quantum simulation. As shown in Table S7, the ground state of H2O molecule can absorb the energy of 150.22 nm UV radiation to generate excited state H2O molecule form that exhibits higher reactivity than the ground state form.

$${{H}_{2}O}\: \underrightarrow{hv\:\left(185\:nm\right)}\:HO\bullet+H\bullet$$
(R-6)

Degradation of PFASs in industrial wastewater via EO + VUV process

 The industrial wastewater sample was collected from the water inlet of industrial wastewater treatment plant in August 2024. Before EO and EO + VUV treatment, wastewater sample was filtered through a 0.22 μm polypropylene membrane to remove suspended solids. Seven out of 25 PFASs were detected in the wastewater sample. The pH, TOC and initial concentration of PFASs of the industrial wastewater sample was shown in Table 3. It is obvious that PFHxA was the largest contributor, which contributed approximately 88.9% PFASs in the wastewater.

Table 3 The pH, TOC and initial concentration of PFASs in industrial wastewater sample.
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The wastewater sample was then subjected to EO and EO + VUV respectively at the current density of 10 mA∙cm−2 for 20 h and the results are shown in Fig. 5. No extra electrolyte addition was needed. Each PFASs in the wastewater sample were removed over 82.7% and 96.2% within 20 h in EO and EO + VUV systems, respectively. PFBA and PFPeA concentration increased in the first four hours in EO and EO + VUV system, due to the degradation of precursors17. After 12-h treatment, 94.8% PFBA, 99.5% PFPeA and 55.8% PFBS were degraded in the EO + VUV system, while 0.0397% PFBA, 88.7% PFPeA and 37.3% PFBS were removed in the EO system, indicating that EO + VUV system exhibited higher degradation efficiency on short-chain perfluoroalkyl acids (PFAAs) than EO system. As shown in Fig. 6, after 2-hr treatment, the removal ratio of PFHxA, HFPO-DA, PFHpA and PFOA were 40.4%, 47.8%, 32.5% and 41.2% in EO system, respectively. However, in the EO + VUV system, the removal rate increased to 55.7%, 82.6%, 85.6% and 81.6% for PFHxA, HFPO-DA, PFHpA and PFOA, respectively. Therefore, the EO + VUV treatment developed herein could be used to remove and degrade PFASs in industrial wastewater at a higher degradation efficiency compared to the EO system.

Fig. 5
Fig. 5
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The concentration profile of PFASs versus time of a treatment with 10 mA∙cm−2 current density (A), or both VUV and 10 mA∙cm−2 current density (B). Error bar represents standard deviations of replicates.

Fig. 6
Fig. 6
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The concentration changes of the seven PFASs after 2-hr treatment in EO and EO + VUV system. Current density: 10 mA∙cm−2. Error bar represents standard deviations of replicates.

Conclusions

The electrochemical degradation combined with VUV irradiation on BDD anode was evaluated for PFOA and PFHxA degradation in this study. It revealed that PFOA and PFHxA degradation was significantly promoted by VUV irradiation. Acidic reaction solution also favored PFOA/PFHxA degradation. UV light, with wavelength under 220 nm, can be absorbed by ground state form of PFOA and PFHxA ions to generate excited state form, which exhibit higher energy and reactivity. The energy (ΔE) required to lose an electron from PFOA and PFHxA ions was reduced, and thus direct electron transfer process was accelerated. Moreover, EO + VUV treatment was also shown higher effectiveness in short-chain PFAAs degradation in real industrial wastewater than EO treatment, reaching 96.2% removal in EO + VUV system compared to 82.2% removal in EO system for PFBA compound. These findings provide guidance for further improvement and optimization of the BDD-based EO reaction systems for PFASs treatment.