All samples including water, sediments, baits, and bullfrogs in different growth periods were collected from typical bullfrog bases around the coast of the South China in June 2021(Figure 1). Water samples were stored in 1-L polypropylene (PP) containers with a narrow mouth and a screw cap, which were collected from downstream and upstream of the water supply river, effluent and influent of the culture ponds in bullfrog base. Corresponding sediment samples were collected with a bottom grab and stored in PP containers from the corresponding locations of water. To reduce individual differences, all mixed biological samples were obtained for tadpoles, juvenile bullfrogs, young bullfrogs, and adult bullfrogs, which were stored in sealed bags and frozen at −20°C in a refrigerator until treatment. All containers and tools were cleaned before use by rinsing sequentially with methanol and Milli-Q water from the sampling site.

All mass-labeled internal standards and native standards were purchased from Wellington Laboratories (Guelph, ON, Canada), including PFBS, PFHxS, PFOS, perfluorodecane sulfonate (PFDS), PFBA, perfluoropentanoic acid (PFPeA), PFHxA, perfluoroheptanoic acid (PFHpA), PFOA, perfluorononanoic acid (PFNA), perfluorodecanoic acid (PFDA), perfluoroundecanoic acid (PFUnDA), perfluorododecanoic acid (PFDoDA), perfluorotridecanoic acid (PFTrDA), perfluorotetradecanoic acid (PFTeDA), perfluorohexadecanoic acid (PFHxDA), and perfluorooctadecanoic acid (PFODA). Four perfluorinated sulfonic acids (PFSAs), 13 perfluorinated carboxylic acids (PFCAs), and chlorinated polyfluorinated ether sulfonate (9Cl-PFONS; F-53B) were measured for all of the samples. 13C4PFBA, 13C4 PFHxA, 13C4 PFOA, 13C4 PFNA, 13C4 PFDA, 13C4 PFUdA, 13C2 PFDoA, 18O2 PFHxS, and 13C4 PFOS were used as internal standards with purity >98%. Methanol (HPLC grade) and acetonitrile were purchased from J.T. Baker (Phillipsburg, NJ, United States). Detailed information is given in Supplementary Table S1.

PFASs analysis was performed using a Thermo Ultimate 3000 Infinity HPLC System equipped with a Thermo TSQ ENDURA LC/MS System, which was operated with electrospray ionization (ESI) in negative ion mode and with multiple reaction monitoring (MRM) of the target analytes. The column used was Agilent ZORBAX Eclipse Plus C18, 2.1 × 100 mm, 3.5 μm (Agilent) connected with a guard column (Agilent). The gradient elution of the mobile phase was 2 mM ammonium acetate (A) and 100% acetonitrile (B). The flow rate was 0.3 ml/min and a 5-μl aliquot of the extracts was injected into the column. Chemical formula, parent ion, quantitative ion, and qualitative ion of PFASs used in study are listed in Supplementary Table S1. Detailed descriptions of instrument conditions are shown in Supplementary Table S2.

To strictly control background pollution, polytetrafluoroethylene or other fluoropolymer materials were prevented from preliminary treatment and analytical processes, and all containers in the laboratory were rinsed with methanol and Milli-Q water before use. There were analyzed in solvent blanks, process blanks, and duplicate samples to make the data reliable. Matrix spikes were based on an internal standard method, which was applied to quantify PFASs shown in Supplementary Table S3. The calibration curve consisted of a series of concentration gradients (0.1, 0.2, 0.5, 1, 2, 5, 10, 20, 50, 100, 150, and 200 μg/L), ensuring that the deviation between the determined value of each point and its theoretical value did not exceed 20%. The limit of detection (LOD) was determined using a signal-to-noise ratio of 3:1 and the limit of quantification (LOQ) was identified as the analyte peak that yielded a signal-to-noise ratio of 10:1. In the statistical analysis, all values lower than LOD were specified as LOD/√2, and in the case of concentrations lower than LOQ, the value of LOQ/√2 was used. The results of LOD and LOQ are shown in Supplementary Table S3.

The distribution coefficient (K_d) is used to analyze the ability of chemical substances adsorbed by the solid phase (sediment) from the liquid phase (water) (Zhou et al., 2013). The field-based K_d was calculated as follows: K d = C sediment / C water where C_sediment is the concentration of PFASs in sediment (ng/g dw), C_water is the concentration of PFASs in water (ng/L), and K_d is in L/g.

Bioaccumulation factors (BAFs) were used to investigate the accumulation of PFASs in bullfrogs, which represents chemical uptake from water. BAF is defined as the ratio of the concentrations of PFASs in biota and concentrations in water (Chevreuil, 2010; Wilkinson et al., 2018; Burkhard, 2021). The field-based BAF was calculated as follows: BAF = C organism / C water

The intake of PFASs from bullfrog was identified as important exposure for human. Therefore, the estimated dietary intake (EDI) of PFASs for local residents was calculated in order to evaluate the human exposure. EDI = ( C bullfrog × C R ) / BW where C is the concentration of PFASs in bullfrog (ng/g dw), C _R is the consumption rate of different pathways (g/day), and BW is the body weight (kg).

The health risks of PFASs for local residents were calculated by the hazard ratios (HR) using the following equation: HR = EDI / RfD where RfD is the reference dose (RfD, ng/kg/day), defined as an acceptable daily exposure without apparent toxicity. RfDs for PFHxA, PFOS, PFOA, PFBS, and PFBA were 320, 2, 3, 430, and 2,900 ng/kg/day, respectively (ANSES, 2017; MDH, 2017). HR higher than 1.0 means a potential risk towards human; however, value under 1.0 means low risk.

Distribution and concentration of PFASs in bullfrog samples during different growth periods were detected from the typical bullfrog bases around the coasts in South China (Figure 3). Total PFASs in all bullfrog tissues were 3.36–84.07 ng/g·dw; PFBA, PFOS, PFOA, PFUdA, PFTrDA, and PFDoA are the predominant PFASs detected in the biota samples. F-53B, which was found at detectable levels in all habitat samples from the aquaculture base, also had a high detection level, reaching 95% with an average concentration of 0.36 ng/g·dw in bullfrog tissues. PFBA, as a short-chain PFAS, was found in all tissues with a concentration of 2.47–48.98 ng/g·dw. Meanwhile, PFBA among all PFASs had the highest concentration in three different particles of baits, which explains why detection level and concentration were high in effluent water at D_eff, F_eff, and H_eff sites. Baits were manufactured by a compressor with a lot of lubricants, possibly containing numerous PFASs, and then biota could ingest contaminated baits or bioaccumulate PFASs via polluted habitats from putting into contaminated baits and industrial pollution (Supplementary Table S5).

PFAS level differences were observed in frogs during different growth periods: adult bullfrog, juvenile bullfrog, young bullfrog, and tadpole. It was evident that sum concentrations of PFASs in adult bullfrog were markedly higher than those in juvenile bullfrog, which could be due to the fact that PFASs had a long-half-life and persistent and bioaccumulative characteristics. Concentrations of PFASs were 184.45 ng/g·dw in the sum of all tested tadpole tissues and 142.65 ng/g·dw in juvenile bullfrog tissues. This result implied the higher total PFAS concentrations observed in tadpoles, which may be responsible for the varying bait between tadpole and juvenile bullfrog. Tadpoles ingested the smaller grain contaminated baits (PFASs: 95.74 ng/g·dw), and juvenile bullfrog ingested the baits of medium grain size (PFASs: 24.77 ng/g·dw). Interestingly, it can be seen that intestines had higher PFAS levels in tadpole than that in juvenile bullfrog on account of undigested contaminated bait in intestines. The concentration of PFOA in tadpole and juvenile bullfrog tissues was 0.01–26.59 ng/g·dw and ND-4.11 ng/g·dw, respectively. As short-chain PFASs, PFHpA (mean concentration: 25.73 ng/g·dw) and PFBA (mean concentration: 6.91 ng/g·dw) had a high detection ratio (100%) and concentration in all tadpole tissues. The current research is not in accordance with previous studies in that PFHxA was the predominant contaminant (80.50 ± 58.31 ng/g and 19.17 ± 12.57 ng/g ww), followed by PFOS (55.02 ± 34.82 and 14.79 ± 6.24 ng/g ww) in both fish from Lake Chaohu, China (Wu et al., 2019).

The distribution in different tissues after uptake largely determines the fate of a chemical and is of great importance for understanding the underlying mechanisms of bioaccumulation. Concentrations of PFASs in samples varied greatly among the different tissues, with the greatest concentrations in the digestive system (intestines and stomach) no matter what stage of growth the bullfrog is in. The concentrations of PFASs in the digestive system (intestines and stomach) were 52%, 45%, 65%, and 52% in adult bullfrog, young bullfrog, juvenile bullfrog, and tadpole, respectively, closely associated with ingesting contaminated bait. It was consistent with other studies in that the highest level in the digestive system among the different tissues was observed in carp and green eel goby, suggesting that PFASs can accumulate by diet from water, sediment, and contaminated bait intake (Hong et al., 2015; Wu et al., 2019). Moreover, liver, as a vital organ for fat metabolism and protein synthesis, also had a higher concentration level of PFASs than muscle. The possible reason for this is that liver could affect the binding of lipid and protein, and the protein content was generally at a higher level in liver while it was at a lower level in muscle (Chevreuil, 2010; Wen et al., 2019; Wu et al., 2019).

Despite the difference in levels, PFASs were found to have the same trend among nearly all muscles of bullfrogs from different growth periods, with PFBA being the most predominant compound in terms of concentration and frequency of detection. The proportion of F-53B in biota samples changed among the different tissues, with the greatest concentrations in liver (0.19–0.74 ng/g·dw). It is worth noting that F-53B was also analyzed from different areas in China, supporting previous findings that this contaminant is widely distributed in the bullfrog tissues from Changshu, Huantai, Zhoushan, and Quzhou (Cui et al., 2018), and consumption of bullfrog may expose humans to PFASs. The scanned diversification among the tissues should be put down to the diverse binding affinity of PFASs for tissue-specific proteins.

Field-based bioaccumulation factors of PFASs in various bullfrog organizations were calculated based on concentrations in water and biota tissues shown in Figure 4. What is noteworthy is that log BAFs of PFTeDA, PFDoA, and PFUdA were obviously higher than those of shorter-chain PFASs including PFBA and PFPeA. The results of this study were generally in agreement with previous results in that the short-chain PFASs are not considered bioaccumulative, but long-chain PFASs with >10 fluorinated carbons were bioaccumulative, potentially due to their large molecular size (Liu et al., 2011b; Hong et al., 2015). This study indicated that the log BAF value was at a higher level in PFOS but at a lower level in PFOA among adult, juvenile, and young bullfrog samples. Analogously, the log BAF value of PFBS was distinctly greater than that of PFBA in the present study, to some extent, which explained that PFSAs could more easily accumulate than PFCAs for similar substances of the same fluorinated carbon chain length.

F-53B, which was supported by international organizations for the production and substitution of PFOS, had a higher mean value of log BAF (with a mean of 0.44) than PFOS (with a mean of 0.31). Interestingly, the half-life of F-53B in humans is longer than that of PFOS in other studies (Loi et al., 2013). Hence, the government and relevant departments should consider F-53B instead of PFOS. Comparing the different ages of bullfrogs, the values of log BAF for long-chain PFCAs (C > 9) were higher in adult bullfrogs than in young bullfrogs; meanwhile, PFSAs, i.e., PFOS, PFBS and F-53B, had a similar trend in that the values of log BAF were higher in adult bullfrogs than in young bullfrogs. The results in this research suggested that long-chain PFCAs and PFSAs had great bioaccumulation propensity and explained why PFSAs and long-chain PFASs had higher concentration levels in adult bullfrogs than in young bullfrogs. The correlation of BAFs between long-chain and short-chain PFASs showed a helpful observation for comprehending the relative significance of diverse bioaccumulation mechanisms.

The values of log BAF in PFASs were 0.38 and 0.13 in juvenile bullfrogs and tadpoles, showing that the log BAFs of individual PFASs varied during different periods of bullfrogs on account of their complex life history following metamorphosis. The mean log BAF of PFOS in liver was 1.18 in adult bullfrogs, which is higher signally than that in muscle, and this was consistent with the trend observed in other studies. In PFOS, the mean log BAF_liver in nine fish species was in the range of 3.5–4.6 from rivers in the Pearl River Delta region, around one log unit higher than in muscle (Pan et al., 2014). The log BAF_liver of PFOS in this research is at a lower level than that found in European chub with log BAF_liver of 4.3 and in rainbow trout with a log BAF_liver of 3.7 (Martin et al., 2003; Labadie and Chevreuil, 2011; Wang et al., 2016b; Liu et al., 2018). This could be explained by the PFAS concentration variability in water, discharge capacity of the surrounding industry, different nutritional habits, and physical and chemical properties of different compounds.

Food dietary intake is the crucial pathway of human exposure to PFASs (Perez et al., 2014). The human exposure risk assessment was performed by calculating the EDI based on the daily consumption of bullfrog and the concentrations of PFASs shown in Figure 5. As body weights and consumption rates vary with age, the EDI of PFASs for five age groups (2–6, 7–12, 13–17, 18–45, and 46–69 years old) was estimated from China (CN), Guangdong province (GD), and Study area (SA), in which the daily consumptions of bullfrog were calculated to be 0.21, 1.16 and 1.89 g/day, respectively (Yan et al., 2015; Huang, et al., 2019; Su et al., 2019).

The EDIs of PFASs via consumption of bullfrog for local residents of CN, GD, and SA are illustrated in Supplementary Tables S6–S8. In general, the total EDI of PFASs ranged from 1.16 × 10⁻² to 4.53 × 10⁻² ng/kg bw/day for residents in China, 6.84 × 10⁻² to 2.26 × 10⁻¹ ng/kg bw/day for residents in Guangdong province, and 1.11 × 10⁻¹ to 3.68 ×10⁻¹ ng/kg bw/day for residents in the Study area. The EDI decreased with age [(2–6) > (7–12) > (13–18) > (18–45) > (45–69)], among which the EDI of residents aged 2–6 years was 3.3 times than that of residents aged 45–69 years, indicating that ingestion of PFASs via bullfrog induced more severe hazards on residents aged 2–6 years than on residents aged 45–69 years. The result was consistent with estimations in most studies (Meng et al., 2019a). It may be relevant that residents aged 45–69 years weighed significantly more than residents aged 2–7 years. The bullfrog breeding scale in Guangdong province is the largest in CN, which accounts for about 40%. The research area is a high-density bullfrog culture area in GD, and the local residents like to consume bullfrog, which leads to the EDI and the intake of bullfrog being the highest among the data.

PFBA, PFUdA, PFTrDA, PFOS, PFOA, PFBS, and F53-B accounted for about 72% of total EDI of the 14 PFASs in CN, GD, and SA. PFBA contributed most to the EDI because of its highest concentration level, followed by PFUdA, PFTrDA, PFOS, PFOA, PFBS, and F53-B. The EDI_CN, EDI_GD, and EDI_SA of PFBA were 3.70 × 10⁻³–1.45 × 10⁻², 2.19 × 10⁻²–7.24 × 10⁻², and 3.55 × 10⁻²–1.18 × 10⁻¹ ng/kg bw/day, respectively. The EDI_2-6(y) of PFBA in SA was 1.18 × 10–1 ng/kg bw/day, which was the highest in all data. EDIs of PFOS and PFOA were also dominant in bullfrog. The exposure of PFOS was 1.07 × 10⁻³–4.20 × 10⁻³ ng/kg bw/day in CN, 6.32 × 10⁻³–2.09 × 10⁻² ng/kg bw/day in GD, and 1.03 × 10⁻²–3.40 × 10–2 ng/kg bw/day in SA, respectively. The EDI_CN, EDI_GD, and EDI_SA of F-53B were 3.93 × 10⁻⁵–1.54 × 10⁻⁴, 2.32 × 10⁻⁴–7.69 × 10⁻⁴, and 3.77 × 10⁻⁴–1.25 × 10⁻³ ng/kg bw/day, respectively. In general, humans were exposed to F53-B in low levels via bullfrog consumption in each age group from all study areas.

The measured EDIs in this study were at lower levels than the values extracted from other studies in China. The mean daily intake via farmed fish in Beijing were 2.40 × 10⁻¹ ng/kg bw/day for PFOS and 4.50 × 10⁻¹ ng/kg bw/day for PFASs (Shi et al., 2012). The EDIs of PFASs through seafood consumption were found in Lianyungang City (59.80 ng/kg/day), in Yancheng City (54.70 ng/kg/day), and in Rizhao City (36.00 ng/kg/day) (Zhang et al., 2019). The Scientific Panel on Contaminants in the Food Chain within the European Food Safety Authority (EFSA) has established tolerable daily intake (TDI) for PFOS and PFOA, i.e., 150 ng/kg body weight/day and 1,500 ng/kg body weight/day, respectively (EFSA, 2008). The mean estimated intakes in the present study were about 1,800 and 2,500 times lower than the TDI for PFOA and PFOS, respectively. Therefore, we can conservatively conclude that there is a low potential exposure risk of PFASs via bullfrog consumption for local residents. However, PFASs were detected in bullfrog samples of different ages, and PFASs are difficult to metabolize and easy to accumulate in the body, which suggested that bullfrog consumption could still bioaccumulate PFASs.

HRs were calculated for PFBA, PFOA, PFBS, PFHxA, and PFOS, and the HRs of PFASs were far below the safety threshold value “1” shown in Supplementary Table S9. The results suggest that frequent consumption of these bullfrogs would not cause severe health effects on residents. Similar results were reported in other studies (Wu et al., 2019; Sun et al., 2021). However, besides the bullfrog, other food species also pose risks for human exposure, which may increase the potential risks for local residents. Nevertheless, the possible health risk related to fluorinate organic compounds should not be overlooked in the future. Hence, added information are needed on the occurrence and exposure of PFASs via aquatic product consumption to assess its long-term risk to human health.