A total of 39 volunteers participated in the study, including 24 women (60%), 14 men (37%), and 1 non-binary (3%), with an average age of 40.7 years (standard deviation (SD) = 10.2 years, range = 26–76 years). Educational level was university or more among 35 (90%) and a high school among 4 (10%). Average consumption of unfiltered tap, bottled, and filtered tap water were, respectively 0.6 (SD = 0.5, range = 0.1–1.5), 0.5 (SD = 0.4, range = 0.3–1.5), and 0.4 (SD = 0.5, range = 0.1–1.5) L/day, based on a self-reported water consumption questionnaire.
PFAS, bisphenol A, and nonylphenol in tap water
In total, 35 PFAS were analyzed in tap water, of which only perfluoroalkyl acids (PFAA; 7 carboxylates and 3 sulfonates) were above the quantification limits, mainly with a carbon chain length shorter than eight (≤C8); while C10, C11 and C12 carboxylates were only detected in one or two samples. Total PFAS detection rate for the first sampling was 79%, and 69% for the second sampling (Table 1). The most frequently detected (>50%) compounds during the first sampling were perfluoropentanoate (PFPeA) (64%; median = 3.3 ng/L), perfluorobutane sulfonate (PFBS) (64%; median = 9.2 ng/L), perfluoroheptanoate (PFHpA) (52%; median = 3.0 ng/L), perfluorohexanoate (PFHxA) (31%; median = 13.0 ng/L) and PFOS (52%; median = 12.5 ng/L), while the other PFAS showed detection frequencies lower than 12% (Table 1, Fig. 1). Similarly, the most prevalent compounds during the second sampling were PFPeA (62%; median = 4.0 ng/L) and PFBS (45%; median = 6.8 ng/L), whereas PFOS and PFHpA were present in 4.8% and 24% samples, respectively (Table 1, Fig. 1). The PFAS composition profile in the first sampling was dominated by PFBS (25.9%), PFOS (22.1%), PFPeA (17.6%), PFHxA (16.2%) relative to the total PFAS concentrations (Fig. 2). In the second sampling, high contributions to total PFAS concentrations were observed for PFPeA (45.7%), and PFBS (39.2%) (Fig. 2). To our knowledge, this was the first study analyzing ether-PFAS (e.g., GenX, and ADONA) in drinking water of the Barcelona region, showing non-detected levels.
Compared to previous studies conducted in Barcelona, replacement PFAS (PFPeA, PFHxA, PFBS) and PFHpA were the most predominant compounds detected in the tap water samples, with observed increasing concentrations over the last 10 years (Supplementary Table 1)16,17.
This dominance of PFAS with fewer than eight carbons (<C8) in drinking water has been confirmed by other studies, following that the fluorochemical industry introduced “short-chain” alternatives to replace the “long-chain” legacy PFAS in formulations26,27. Initially, replacement PFAS were assumed safer given their lower bioaccumulative potential relative to legacy PFAS, however recent studies raised concerns about their high persistence, high mobility in the aquatic environment and adverse human health effects26,27,28,29. Notably, these compounds are less hydrophobic, having stronger polarity and lower adsorption potential to soil, which makes them more mobile and allows them to penetrate to deeper ground layers of water30. In addition, studies showed that conventional source water treatment technologies using activated carbon have been less effective to remove replacement PFAS31,32.
For legacy compounds such as PFOA, perfluorononanoate (PFNA) and PFOS, a decreasing trend in concentrations in tap water was observed over the last 10 years in Barcelona compared to previous studies (Supplementary Table 1)16,17. PFAS composition profiles suggest that legacy compounds (PFOS, PFOA) still contribute to total PFAS concentrations, although not consistently across sampling events (Fig. 2). This may be explained by the high persistence and accumulation of legacy PFAS in the environment that can lead to human exposure long after being discontinued in global production9.
Our results showed that median total PFAS concentrations were three times higher during the first sampling (30.0 ng/L) compared to the second (9.8 ng/L) (Table 1, Fig. 1). The second sampling was conducted after the rainy season in May whereas the first sampling was conducted in August-September during late-summer months. Differences in PFAS concentrations across sampling campaigns may be due to the seasonal variation of the quality of surface waters that supply the drinking water for Barcelona (Ter and Llobregat rivers). Indeed, a study conducted in Catalonia (Spain)33 found seasonal variation over 3 sampling campaigns of untreated water from the Ebro river for PFPeA (autumn = 30%, winter = 17%, spring/summer = 66%) and PFOS (autumn = 22%, winter = 4%, spring/summer = 86%). Seasonal changes in PFAS concentrations of surface and groundwater have been observed in different countries34,35. For instance, Nguyen et al. (2022) investigated the catchment of a river in Sweden at a sampling site impacted by the use of PFAS-containing aqueous fire-fighting foams (AFFFs), where they found higher PFAS concentrations due to the high water flow season (i.e. spring). Additionally, they also found inverse seasonal trends in PFAS concentrations at sampling sites that were less impacted by point sources, that can possibly be explained by the effect of dilution during high flow events without extra inputs of pollution34. In another study, Tokranov et al., (2021) found lower concentrations of PFAS in the summer and higher concentrations during the winter months within the surface water/groundwater boundary and in downgradient groundwater of a lake (Massachusetts, USA) driven by natural biogeochemical fluctuations associated with surface water/groundwater boundaries. Taken together, seasonal differences in PFAS levels of source water have been documented with different underlying mechanisms, therefore it is important to note that seasonal changes may have an influence on drinking water quality.
The results of spearman correlation coefficients between individual PFAS are summarized in Fig. 3. The correlations did not reach statistical significance, however strong positive correlations were observed between PFBS and Total PFAS (r = 0.6; p value = 0.4) and moderate correlations between PFPeA and Total PFAS (r = 0.4; p value = 0.7); PFPeA and PFBS (r = 0.4; p value = 0.8); PFOS and Total PFAS (r = 0.5; p value = 0.7) in the first sampling. Regarding the second sampling, PFPeA was highly correlated with Total PFAS ( = 0.8; p value = 0.4) and moderately with PFBS (r = 0.5; p value = 0.4). Our results are in line with a previous study showing moderate or high correlations between individual PFAS in treated water that have been explained by their similar sources36. A limitation of the correlation analysis was the number of samples above the limit of quantification only a few compounds were included for this analysis for the compounds.
Policies to manage PFAS contamination are being implemented at EU level, including the recent EU DWD24, that regulates PFAS as a class to be routinely monitored in drinking water starting in 2023. In this study, the sum and total PFAS concentrations as defined by the EU DWD were identical as only carboxylates and sulfonates C4-C12 were detected. The observed median sum/total of PFAS concentration in the first sampling (30.0 ng/L) was lower than the EU DWD regulatory limits, except for one sample (180 ng/L) that exceeded the parametric value for the “sum of PFAS” (Fig. 4). In this sample, PFPeA, PFBS, PFOA, PFOS, PFHpA and PFHxA were quantified at concentrations of 72, 52, 21, 13, 12, and 10 ng/L, respectively, and the sum of carboxylates represents 64% of total concentration level. The corresponding sum/total PFAS concentration of the second sampling (7.6 ng/L) was considerably lower and below the EU parametric value (Fig. 4).
Bisphenol A and nonylphenol were not detected in tap water samples (Table 1).
Results are representative of urban settings supplied with surface water with diffuse source contamination by PFAS and other industrial chemicals.
PFAS, bisphenol A, and nonylphenol in filtered and bottled water
In this study, the activated carbon (AC) pitcher filtered samples showed similar PFAS levels as the respective tap water samples before filtering. Median concentrations were 32.0 ng/L in non-filtered, and 33.0 ng/L after AC filtration (Table 2). Some samples showed slightly higher PFAS concentrations after AC filtering. Given that the AC filters remove contaminants through adsorption process, we hypothesize that domestic AC filters in real-life working conditions do not efficiently adsorb PFAS when highly loaded and clogged, and thus having the potential to release PFAS to the filtered water. PFAS breakthrough of AC filter has been observed when AC media was not regenerated to renew adsorptive capabilities37. Flores et al. (2013) previously showed the importance of the loading of granular AC filters to guarantee the efficient removal of PFAS in drinking water potabilization processes. On the other hand, removal efficiency of PFAS by AC pitcher filters from drinking water was recently evaluated by Herkert et al., (2020) reporting that 85% of activated carbon filters significantly removed PFAS by ~50% in drinking water, with increased removal efficiency for legacy PFAS. Overall, evidence shows that carbon filters can effectively remove PFAS, only if properly maintained37.
Reverse osmosis (RO) technology for water treatment has been effective to remove contaminants by pushing water through a semipermeable membrane and provide consistent removal for longer period (6–12 months)37. Our results show that RO filters reduced median PFAS concentrations from 38.0 to 1.0 ng/L (97% reduction) (Table 2). Consistently, previous studies showed that domestic RO filters removed more than 90% of PFAS from tap water38, as well as effectively removed PFAS during potabilization process at a drinking water plant32. Particularly, RO filters have been proven for their ability to remove both replacement and legacy PFAS to below detection limits, due to the membrane performance attributable to the small size of pores37.
In the current study, PFAS were not detected in bottled water, consistently with previous studies that did not detect PFAS in 4 Spanish bottled water brands17 and in 20 Japanese and international bottled water brands10. On the contrary, Ericson et al. (2008) found low PFAS concentrations (<1 ng/L) in three Spanish bottled water brands, and Schwanz et al. (2016) quantified PFAS in 10 Spanish bottled water brands with median concentrations of 11 ng/L for the sum of PFAS. Moreover, other studies reported the occurrence of PFAS (slightly above LOQ) in mineral water samples from Europe11,20 and from the United States39.
Bisphenol A and nonylphenol were not detected in filtered and bottled water samples.
Occurrence of PFAS in urine
We detected 5 distinct PFAS, one each in 5 out of 39 urine samples (Fig. 2): PFPeA (0.018 ng/mL; 65.9 ng/g creatinine), PFHxA (0.013 ng/mL; 325 ng/g creatinine), 6:2 FTS (0.024 ng/mL; 11.6 ng/g creatinine), 8:2 FTS (0.009 ng/mL; 4.3 ng/g creatinine) and PFOSA (0.026 ng/mL; 10.8 ng/g creatinine). A previous study conducted in Barcelona (N = 30) found 8 PFAS in urine (PFBA, PFHxA, PFHpA, PFOA, PFDA, PFUdA, PFBS, PFHxS) above detection limits, of which PFBA (median = 337.0 ng/mL) was detected in 100% of the samples40. Recent studies are in line with our results showing that PFAS are detected in urine at lower concentrations compared to urinary concentrations detected for communities that work or live close to polluted sites or occupationally exposed41,42,43. Calafat et al. (2019) showed that 67.5% of the US general population did not have detectable urinary PFAS concentrations. In the current study, three of the detected PFAS in urine samples (6:2 FTS, 8:2 FTS, PFOSA) have not been present in drinking water. Two participants with detectable 6:2 FTS or PFOSA concentrations reported consuming bottled water (where PFAS were undetected), and three out of five urine samples with detectable levels of PFPeA, PFHxA, or 8:2 FTS were from participants reporting AC filtered water consumption. Altogether, findings suggest that drinking water might be responsible for PFPeA and PFHxA urinary levels, while exposure sources other than drinking water explain the urine concentrations of 6:2 FTS, 8:2 FTS, and PFOSA. Our results are consistent with Zhang et al. (2013) regarding the detection of replacement PFAS in urine (PFPeA, PFHxA) that have shorter half-lives in humans, thus urine is considered a suitable biospecimen for PFAS that are rapidly cleared from the human body44. On the other hand, PFAS can bind to blood protein and the body burden is reflected by serum levels of PFAS that can affect the transfer efficiency to urine45. A limitation of the present study is that it involved spot urine samples instead of a repeated sampling method.
Non-target screening in drinking water
A summary of tentative results regarding non-target screening is shown in Table 3. A total of 16 out of 248 analytes were detected in at least one water sample, with occurrence frequency varying substantially between types of drinking water (Table 3, Supplementary Table 2). Non-filtered tap water presented the highest number of compounds including 12 micropollutants and 4 metabolites. The highest detection rates were found for carbamazepine (a recalcitrant pharmaceutical compound used as anticonvulsant and mood stabilizer), tris(chloroisopropyl) phosphate (a high-volume production chemical included in polymer formulations because of its flame retardant potential), suberic, and azelaic acids (saturated linear dicarboxylic acids used in plastic manufacturing personal care products), and terbuthylazine (herbicide). These were detected, respectively, in 100%, 83%, 78%, 56%, and 56% of unfiltered tap water samples (Table 3).
Tap water filtered with AC showed lower detection frequency relative to unfiltered tap water, but more than twice compared as RO filtered water. Notably, personal care products (suberic and azelaic acids) were detected in all types of drinking water (non-filtered tap, filtered tap, and bottled). However, interpretation of findings should be cautious. According to the confidence levels of non-target analysis in high resolution mass spectrometric analysis we could confirm 4 out of 5 levels, i.e., level 2 according to Schymanski et al. (2014), specifically: (a) the mass of interest; (b) the unequivocal molecular formula but insufficient structural evidence; (c) the tentative candidate compounds by identifying the suspect, substructure, and class; (d) the probable structure by library diagnostic evidence46. In this study, we could not confirm probable structure by a reference standard and validate the results nor could we quantify the concentrations of suspects. Hence, we only reported instrumental response (a.u = arbitrary units) or, in other words, the presence and frequency of suspect contaminants.
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