Human Exposure Pathways to Per- and Polyfluoroalkyl Substances (PFASs)-A Comprehensive Review of Sources, Physicochemical Properties, and Human Health Risk Assessment.

The history of PFASs is rooted in the rapid advancement of organofluorine chemistry during the first half of the 20th century. A pivotal moment occurred in 1938 when Dr. Roy Plunkett, working for DuPont, discovered polytetrafluoroethylene (PTFE) while researching refrigerant gases. This material, branded as Teflon, exhibited unprecedented thermal resistance, low friction, and chemical inertness. Its utility was proven during World War II within the Manhattan Project, which catalyzed the industrial expansion of fluoropolymer technologies [ 6 ]. Commercialization intensified in the 1950s, leading to the synthesis of other PFASs such as perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA). Their unique ability to repel both water and oil made them indispensable for surface treatments and fire suppression [ 6 , 7 ]. However, the same properties that ensured industrial success led to global environmental persistence. After 2000, the recognition of their bioaccumulative nature and health risks led to the tightening of regulations, particularly for long-chain PFASs, and the formalization of definitions by organizations like the OECD [ 6 , 7 ]. The 2021 OECD definition classifies PFASs as organic compounds containing at least one fully fluorinated methyl (–CF3) or methylene (–CF2–) group. This updated definition is more inclusive than the 2011 Buck et al. criteria, which focused primarily on perfluoroalkyl moieties (C n F2 n+1 –) [ 1 , 8 ]. This shift reflects the complexity of the PFAS landscape; while a 2018 report cataloged 4730 PFASs, the 2021 criteria expanded this number to over 7 million potential substances in the PubChem database. Such diversity underscores the need for standardized nomenclature to facilitate effective monitoring and regulation [ 1 , 8 , 9 ]. The primary distinction lies between perfluoroalkyl substances (where all C–H bonds in the alkyl chain are replaced by C–F bonds) and polifluoroalkyl substances (where at least one C–H bond remains). Polyfluoroalkyl compounds often serve as precursors, degrading in the environment into highly stable perfluoroalkyl “terminal” products [ 1 ]. PFASs are further categorized by functional groups, such as perfluoroalkyl carboxylic acids (PFCAs, e.g., PFOA) and perfluoroalkyl sulfonic acids (PFSAs, e.g., PFOS). Furthermore, chain length is a critical factor: long-chain PFASs (PFCAs ≥ C7, PFSAs ≥ C6) are more bioaccumulative, while short-chain alternatives, introduced as safer substitutes, exhibit higher environmental mobility and different toxicological profiles [ 1 , 4 , 10 ]. The unique properties of PFASs determine both their industrial utility and the difficulty of environmental remediation. The C–F bond, one of the strongest in organic chemistry (~485 kJ/mol), provides resistance to hydrolysis, photolysis, and biodegradation, earning them the name “Forever Chemicals” [ 1 , 11 , 12 ]. Their amphiphilic nature allows them to repel both water and oil, making them superior surfactants compared to hydrocarbon-based alternatives [ 2 , 4 , 13 ]. Solubility is inversely proportional to chain length, dictating their distribution between aqueous phases and solid matrices like soil or sediment [ 14 ]. The specific physicochemical parameters of the most widely studied PFASs, namely PFOA and PFOS, are summarized in Table 1 . PFASs resist natural degradation, leading to continuous accumulation since the 1950s [ 10 , 14 ]. Short-chain compounds are highly mobile in aqueous systems, while long-chain compounds adsorb strongly to soil particles and bottom sediments [ 10 , 14 ]. Both undergo long-range atmospheric and oceanic transport, reaching remote indicator areas like Antarctica, where they are detected even in penguins [ 3 , 10 ]. Unlike classic POPs that accumulate in lipids, PFASs bind to transport proteins such as serum albumin and fatty acid-binding proteins (FABPs), concentrating in the liver, kidneys, and blood [ 15 , 16 ]. They exhibit trophic magnification, with levels increasing toward apex predators [ 17 ]. Toxicological effects include PPAR activation, oxidative stress, and endocrine disruption, with long-chain PFASs exhibiting potential carcinogenic effects, as well as reproductive and developmental toxicity [ 4 , 10 , 18 , 19 , 20 ]. Human elimination half-lives are exceptionally long: ~3.8 years for PFOA, 5.4 years for PFOS, and up to 8.5 years for PFHxS [ 4 , 10 ].

Widespread global PFAS production and use in over 200 OECD categories explain their ubiquity, with historical emissions of long-chain PFCAs estimated at 2610–21,400 tons [ 13 , 21 , 22 ]. Primary emissions stem from AFFFs at military and civilian airports, firefighting training centers, and petrochemical facilities (such as oil refineries) as well as industrial discharges from fluorochemical, textile, and paper plants [ 4 , 13 , 23 , 24 , 25 ]. Consumer exposure occurs via migration from food contact materials (FCMs), non-stick cookware (particularly when overheated >260 °C) [ 12 ], and indoor dust [ 26 , 27 ]. Protective masks also constitute a relevant consumer product source of exposure [ 28 ]. The primary sectors of PFAS application and their corresponding human exposure routes are summarized in Table 2 . PFASs contaminate groundwater via point sources, such as AFFF training sites, and diffuse sources, including the application of biosolids on agricultural land [ 25 ]. In the US, PFASs are detected in ~45% of tap water samples [ 33 ]. In Europe, approximately 23,000 contaminated sites have been identified, including 2300 hotspots [ 34 , 35 ]. Conventional water treatment is ineffective; advanced remediation (GAC, reverse osmosis) is required but shows diminishing efficacy for short-chain compounds [ 14 , 36 ]. Bottled water does not automatically represent a safe alternative, as PFASs can also be present in these products [ 36 , 37 ]. Global regulatory bodies have established varying thresholds for PFASs in drinking water to mitigate these risks [ 35 , 37 , 38 , 39 , 40 ], as compared in Table 3 . Dietary intake via aquatic, plant, and animal-derived products constitutes a major human exposure pathway [ 44 , 45 ]. Ingestion of contaminated seafood is highly significant due to protein binding in fish liver and serum [ 15 , 16 ], with PFOS exhibiting a higher trophic magnification factor (TMF) than PFOA [ 17 ]. Migration of up to 68 different PFASs from paper and cardboard packaging also contributes significantly, influenced by contact time and fat content [ 29 , 30 ]. Furthermore, PFAS accumulation in crops (e.g., leafy vegetables, strawberries) grown in biosolid-treated soil [ 5 , 25 , 44 ], and their subsequent transfer to meat, milk, and eggs, integrates these chemicals deeply into the human food chain [ 46 ]. Indoor dust serves as a primary reservoir for PFASs migrating from treated furniture, building materials, and carpets [ 26 ]. Profiles vary by building age; older buildings are dominated by PFOS, while newer ones show short-chain PFASs like PFBS [ 26 , 27 ]. Children are at higher risk due to hand-to-mouth behavior and proximity to floors, often exhibiting higher serum levels than their mothers [ 47 ]. Industrial emissions, such as the estimated 0.4–1.3 tons of PFOA emitted annually from a Chinese industrial park, facilitate regional and global transport at altitudes above 1500 m [ 48 , 49 ]. This atmospheric transport drives contamination in remote regions like the Arctic (alert station, 82° N), where monitoring (2014–2023) shows resurgent levels of short-chain PFASs like TFA and PFBA [ 25 , 48 , 49 , 50 ]. Dermal exposure occurs primarily through cosmetics and DWR-treated clothing [ 31 , 32 ]. In a study of 231 cosmetics from the US and Canada, over half contained high fluorine levels, yet under 8% listed PFASs on labels; concentrations in highly fluorinated products ranged from 22 to 10,500 ng/g [ 31 ]. Estimated dermal exposure from cosmetics can reach a median of 1.28 ng/kg bw/day, which may independently exceed EFSA’s recommended limits [ 51 ]. Direct, prolonged contact with DWR-treated outdoor textiles provides another consistent exposure route [ 32 ]. While initial assessments frequently focused on the deposition of PFASs on the skin surface, recent evidence confirms that these compounds actively penetrate through the dermal barrier to enter the systemic circulation. The transdermal permeation efficiency of PFASs is heavily dictated by their physicochemical properties, particularly the carbon chain length and lipophilicity. Short-chain PFASs demonstrate substantial dermal bioavailability. In in vitro 3D human skin equivalent (3D-HSE) models exposed for 36 h, the absorbed fractions of perfluoropentanoic acid (PFPeA) and perfluorobutane sulfonate (PFBS) reached 58.9% and 48.7%, respectively [ 52 ]. Similarly, when human skin models were exposed to PFAS mixtures for 24 h, only short-chain compounds (such as FBSA, PFBA, PFPrS, and PFPeA) successfully penetrated into the receptor fluid [ 53 ]. Conversely, dermal permeation decreases significantly as the carbon chain length increases. For the legacy long-chain compound perfluorooctanoic acid (PFOA), the absorbed fraction in 3D-HSE models was 13.5% after 36 h [ 52 ], whereas other rigorous evaluations recommend a more conservative fractional absorption value of 1.6% for finite dose human exposure assessments [ 54 ]. Crucially, the absence of immediate systemic permeation for long-chain PFASs does not eliminate their toxicological risk, primarily due to the skin reservoir effect. Highly lipophilic, long-chain PFASs (Cn ≥ 9) exhibit a strong affinity for skin tissue and accumulate within the epidermal layers. Studies have shown that up to 66.5% of applied perfluoroundecanoic acid (PFUnDA) and 68.3% of perfluorononane sulfonate (PFNS) are retained directly within the skin tissue, with no immediate detection in the absorbed fluid fraction [ 52 ]. Ex vivo diffusion studies on porcine skin also indicate that a majority of applied PFOA and PFBA is retained in the upper skin layers rather than permeating entirely [ 55 ]. This tissue retention creates a localized reservoir from which PFASs are continuously and slowly released into the systemic circulation. In in vivo rat models, the peak times of PFASs in systemic circulation following dermal exposure were found to be between 8 and 72 h—markedly longer than the 1 to 24 h observed following oral administration—which highlights the prolonged nature of dermal penetration [ 56 ]. Once absorbed systemically, dermally applied PFASs induce significant toxicological outcomes. Subchronic topical exposure to PFHpS and PFOS in murine models successfully elevated serum and urine concentrations, triggering hepatotoxicity (increased relative liver weight) and marked immunosuppression, including reduced B-cells and macrophages [ 57 ]. Sublethal developmental effects from dermal contact have also been documented in postmetamorphic amphibians at lowest-observable-effect concentrations (LOEC) of 50 to 120 ppb [ 58 ]. Additionally, emerging predictive models suggest that neutral PFASs can undergo direct transdermal absorption from the gas phase, potentially making indoor air a more significant dermal exposure pathway than inhalation for certain compounds [ 59 ]. To systematize the kinetic variables driving transdermal penetration and tissue retention, key absorption parameters for selected PFASs are summarized in Table 4 . Human exposure to PFASs exhibits significant regional and demographic variability, revealing profound disparities between the general public and highly exposed cohorts. Large-scale biomonitoring in the United States (NHANES 1999–2018) demonstrates that factors such as age, sex, and race or ethnicity strongly influence serum PFAS concentrations, with adult males and specific racial groups exhibiting higher exposure risks [ 60 ]. Furthermore, spatial mapping of pregnant women in the US indicates that geographic location and proximity to local point sources can often overshadow expected demographic trends, highlighting critical environmental justice considerations regarding vulnerable populations [ 61 ]. In Europe, the HBM4EU biomonitoring initiative revealed that 14.3% of teenagers overall, and up to 24% in Northern and Western Europe, possessed serum PFAS levels exceeding the health-based guidance equivalent of 6.9 μg/L established by EFSA [ 62 ]. These background population burdens contrast sharply with the extreme concentrations identified in environmental hotspots and occupational environments. For instance, the HBM4EU study recorded 95th percentile serum levels reaching an alarming 192 μg/L among chrome plating workers [ 62 ]. Consequently, the highest documented PFAS levels occur in professional settings [ 6 , 23 , 63 ]. Firefighters face severe chronic exposure from AFFFs and protective gear, showing significantly elevated serum levels [ 6 , 32 , 63 , 64 ]. Workers in fluorochemical manufacturing [ 63 , 65 ] and chrome plating (where PFOS acts as a mist suppressant) [ 6 , 23 , 63 ] are exposed primarily via the inhalation of aerosols. Additionally, professional skiers and ski equipment technicians are at risk through the inhalation of fluorinated wax fumes [ 63 ], and textile industry workers face exposure during fabric impregnation processes [ 63 , 65 ].

Increasing human exposure underscores the urgent need to reassess not only legacy compounds, but also emerging alternatives that exhibit significant bioaccumulative properties [ 66 ]. Evidence mapping across more than 150 PFAS compounds highlights substantial gaps in current toxicological understanding and statistical modeling approaches, particularly regarding systemic human health effects [ 67 ]. Consequently, contemporary clinical and epidemiological assessments must move beyond reductionist, linear models and adopt a more holistic interpretation of toxicological and population-based data. When interpreting these findings, a fundamental methodological distinction must be made between statistical association and biological causality. The majority of the human epidemiological data discussed herein derive from observational cross-sectional or case–control studies. While these robustly identify associations between elevated PFAS serum concentrations and adverse health outcomes, they are inherently limited in establishing definitive causal pathways due to potential unmeasured confounding factors, co-exposures, and the possibility of reverse causality. Early epidemiological studies contributed to a strong narrative regarding PFAS carcinogenicity; however, this paradigm is currently undergoing substantial revision. The landmark C8 Science Panel reported “probable links” between PFOA exposure and kidney cancer as well as thyroid disease, has been critically reassessed in light of improved exposure classification, confounder control, and methodological rigor. Recent analyses fail to confirm a consistent causal relationship in general populations [ 68 ]. Similarly, evidence linking PFASs to thyroid cancer remains inconclusive, with meta-analyses demonstrating high heterogeneity and inconsistent associations ranging from positive to null or even inverse relationships [ 69 ]. Notably, specific high-exposure cohorts, particularly in China, report associations between PFAS mixtures and increased risks of nodular goiter and papillary thyroid carcinoma [ 70 ]. A marked discrepancy exists between experimental and epidemiological findings regarding liver and gastrointestinal cancers. In vivo and in vitro studies consistently demonstrate that PFASs can induce hepatocellular proliferation and tumor formation at high doses [ 71 ]. However, human epidemiological evidence remains inconsistent. Studies on colorectal cancer report mixed outcomes, including harmful, null, and even potentially protective associations under specific conditions [ 72 ]. Prenatal exposure studies suggest limited and compound-specific effects on testicular cancer risk, with sulfonates associated with slightly increased risk and carboxylates with decreased risk [ 73 ]. Mechanistically, PFAS-related carcinogenesis appears to be largely indirect, mediated through disruptions in lipid metabolism, oxidative stress, and activation of pro-oncogenic pathways [ 74 ]. In contrast to oncological inconsistencies, evidence for PFAS-induced organ toxicity is comparatively robust. Short-chain PFASs have been associated with elevated alanine aminotransferase levels, indicating early hepatic injury [ 75 ]. Multiple studies link PFAS exposure to the progression of non-alcoholic fatty liver disease (NAFLD), with sex-dependent differences showing greater susceptibility among women [ 76 ]. Both population-based and experimental studies confirm intrahepatic toxicity and steatosis associated with PFOA accumulation [ 77 ]. In pregnant women, PFAS exposure modulates liver function parameters, reflecting cumulative toxic effects [ 78 ]. Mixture exposure analyses reveal systemic biomarker perturbations, suggesting multi-organ suppression [ 79 ]. Advanced modeling approaches further indicate that combined exposure to PFASs and toxic metals contributes to chronic kidney disease through sustained oxidative stress mechanisms [ 80 ]. Cardiometabolic effects of PFASs are highly dependent on age, sex, and chemical structure. Contrary to earlier assumptions, many cross-sectional meta-analyses do not show a consistent association between major PFAS isomers and metabolic syndrome in the general population [ 81 ]. However, compound-specific associations have been observed: PFNA correlates with increased risk in adolescents, whereas PFHxS shows inverse associations in young adults [ 82 ]. Mixture exposure has been linked to metabolic syndrome predominantly in women [ 83 ]. In older populations, long-chain PFASs are associated with dysglycemia, reduced HDL levels, and increased waist circumference [ 84 ]. PFAS exposure also alters lipid metabolism in non-occupationally exposed populations [ 85 ], including associations with LDL cholesterol [ 86 ], thereby contributing to cardiovascular risk via lipid-mediated pathways [ 87 ]. Notably, dose–response relationships for type 2 diabetes often exhibit non-linear (parabolic) patterns [ 88 ], while uric acid mediates increased gout risk under inflammatory conditions [ 89 ]. Early-life exposure to PFASs influences adipocyte development and body composition trajectories. Alterations in adiposity have been observed in children, with clear sex-specific differences: prenatal exposure to PFAS mixtures is associated with decreased adiposity in boys but increased fat accumulation in girls [ 90 , 91 ]. In infancy, adiposity patterns reflected in weight-for-age z-scores (WAZ) are influenced by maternal PFAS burden [ 92 ]. Metabolomic studies further demonstrate disruptions in amino acid pathways linked to adipogenesis [ 93 ], confirming that metabolic dysregulation originates during fetal development [ 94 ]. The placenta represents a critical target of PFAS toxicity, mediating the relationship between prenatal exposure and later health outcomes [ 95 ]. Maternal exposure has been associated with accelerated epigenetic aging in cord blood [ 96 ]. Long-chain PFASs may reduce birth weight and alter placental hemodynamics [ 97 ], although these effects are modified by co-exposure to toxic metals [ 98 ]. Additional mechanisms include altered adipocytokine expression [ 99 ] and changes in anogenital distance [ 100 ]. Increased risks of congenital anomalies [ 101 , 102 ] and preterm birth [ 103 ] have also been reported. PFAS exposure disrupts maternal sleep architecture, contributing to systemic oxidative stress [ 104 , 105 ], and is implicated in gestational diabetes [ 106 ], particularly under modifying conditions such as infectious disease exposure [ 107 ]. Thyroid axis dysregulation within the maternal–fetal unit is well documented [ 108 , 109 ], while associations with hypertensive disorders of pregnancy remain heterogeneous across populations [ 110 , 111 ]. PFAS exposure affects reproductive function in both sexes, though clinical manifestations vary. In women, PFASs are associated with altered ovarian function and increased risk of infertility and early menopause [ 112 ]. In vitro studies demonstrate receptor dysregulation in human granulosa cells [ 113 ], while uterine smooth muscle exhibits increased proliferative activity [ 114 ]. Clinically, higher PFOS levels correlate with increased prevalence of polycystic ovary syndrome [ 115 ]. In assisted reproduction, PFAS concentrations in follicular fluid are associated with reduced embryo quality and impaired cleavage rates [ 116 , 117 ]. In men, PFAS exposure is linked to decreased semen quality [ 118 ], suppression of Leydig cell markers and estradiol levels [ 119 ], and disruption of the blood–testis barrier [ 120 ]. Paternal exposure may also induce epigenetic alterations in sperm, affecting offspring transcriptomic profiles [ 121 ], with combined hepatotoxicity further exacerbating reproductive damage [ 122 ]. The neurotoxic effects of PFASs are characterized by substantial heterogeneity and analytical complexity. Some cohort studies report associations between PFOA exposure and increased risks of autism spectrum disorder (ASD), attention-deficit/hyperactivity disorder (ADHD), and developmental delays, often following non-linear dose–response relationships [ 123 , 124 , 125 ]. Conversely, inverse associations observed for PFAS mixtures are likely attributable to systematic bias rather than true neuroprotective effects [ 125 ]. Prenatal exposure has been linked to reduced cognitive performance in early childhood [ 126 ], although some studies report improved language outcomes and faster reaction times in adolescents [ 127 ]. Sleep disturbances may originate in utero, manifesting as increased nocturnal awakenings in infancy [ 128 ]. In adults, sex-specific patterns emerge, with inverse associations between PFAS exposure and anxiety in men, but strong positive associations with anxiety and depression in women [ 129 ]. PFAS exposure induces immunosuppression, particularly at epithelial barriers such as the skin and mucosa [ 130 ]. Impaired adaptive immune responses and reduced vaccine efficacy, including diminished tetanus antibody production, represent critical public health concerns in exposed children [ 131 ]. These effects are accompanied by systemic inflammation and cytokine dysregulation mediated by macrophage activation [ 132 ]. While inhalation exposure increases asthma risk [ 133 ], paradoxical reductions in allergic symptoms have been observed in some infants, potentially due to prenatal immune modulation [ 134 ]. Additional systemic effects include telomere length alterations [ 135 ] and interference with vitamin D and folate metabolism [ 136 ]. Bone tissue responses are heterogeneous, with PFOS associated with decreased bone mineral density, whereas PFBS exposure has been linked to increased lumbar spine density [ 137 , 138 ]. Dental mineralization is similarly compound-specific [ 139 ]. Finally, PFAS exposure significantly alters the gut microbiome, reducing beneficial bacterial populations and increasing pathogenic taxa from early infancy [ 140 ]. Given the breadth and heterogeneity of PFAS-associated health outcomes across multiple organ systems, a structured synthesis of epidemiological and experimental evidence is presented in Table 5 . This table integrates findings across metabolic, hepatic, endocrine, reproductive, neurodevelopmental, and immunological domains, highlighting both consistent patterns (e.g., hepatotoxicity, immunosuppression) and areas of substantial uncertainty (e.g., carcinogenicity, neurobehavioral outcomes). A critical limitation in current PFAS epidemiology is the inherent risk of exposure misclassification. Many studies rely on single temporal spot samples (e.g., one-time blood draws) to estimate historical or chronic exposure, which may fail to accurately capture the cumulative toxicological burden. Furthermore, the interpretation of biomonitoring data is deeply confounded by the extensive variability in biomarkers. The biological half-lives of PFASs exhibit significant inter- and intra-individual differences influenced by genetics, age, and sex-specific excretion pathways—such as menstruation, lactation, and transplacental transfer—complicating accurate dose–response modeling.

Polar regions, due to their remote location and absence of local industrial emission sources and relatively pristine ecosystems, function as natural laboratories for observing the long-range transport of pollutants through the atmosphere and oceans [ 143 , 144 ]. The very presence of PFASs in Antarctic environments serves as direct evidence of their global distribution and provides an early warning of intensification of pollution, consistent with the One Health approach linking environmental, animal, and human health perspectives [ 3 ]. Indicator organisms such as penguins of the genus Pygoscelis allow for effective monitoring of polar ecosystem contamination through the analysis of PFAS presence in eggshells, feathers and guano [ 3 ]. Temporal trend analysis proves particularly valuable in this regard: a study of Adélie penguin ( Pygoscelis adeliae ) eggs spanning 1997–2021 demonstrated a decline in PFOS concentrations consistent with the timeline of international bans on its use, alongside a concurrent increase in long-chain PFCA concentrations, a group not yet subject to regulation during the study period [ 145 ]. These findings confirm the utility of polar environments as a tool for evaluating the effectiveness of global PFAS regulations, while simultaneously highlighting the need to update existing legal frameworks to address emerging contaminants and close the regulatory gaps that allow unregulated substitutes to perpetuate environmental contamination on a global scale.

Per- and polyfluoroalkyl substances (PFASs) constitute a vast group of thousands of synthetic organic chemicals characterized by the presence of at least one fully fluorinated carbon atom. This structural feature, where hydrogen atoms bonded to carbon are replaced by fluorine, defines the unique chemical identity of this group [ 1 ]. PFASs have been integrated into various industrial sectors for decades, valued for their exceptional thermal stability and surfactant properties. Since the 1940s, they have been utilized in the production of non-stick coatings for packaging and cookware, protective clothing, aqueous film-forming foams (AFFFs), electronics, semiconductors, and numerous other consumer products [ 1 , 2 ]. The most distinctive feature of PFASs is their extraordinary stability, stemming from the high-energy C–F bonds. These bonds provide chemical and thermal resistance, rendering the molecules virtually immune to both biotic and abiotic degradation in the environment. Consequently, mitigating PFAS pollution is a complex challenge exacerbated by a significant lack of effective large-scale neutralization methods [ 1 ]. Growing environmental evidence confirms the ubiquitous presence of PFASs in natural ecosystems and their ability to migrate across environmental media, persisting for decades. Studies have documented PFASs in wildlife, surface waters, sediments, and the atmosphere. Notably, contamination is found even in remote regions lacking local emission sources, proving the capacity of these compounds for long-range transport [ 1 , 3 ]. Current toxicological research correlates PFAS exposure with significant health risks, including increased incidences of cancer, infertility, liver and kidney diseases, and disruptions of the cardiovascular and immune systems [ 4 ]. Humans are exposed to PFASs through occupational and lifestyle pathways, including inhalation, ingestion (contaminated water and food), and dermal contact with consumer products [ 2 ]. Due to their persistence and bioaccumulative nature, PFASs represent a critical public health challenge, prompting a surge in scientific interest and an increasingly stringent global regulatory landscape [ 1 , 5 ]. This review analyzes current knowledge regarding exposure sources and evaluates regulatory efficacy based on global pollution trends. Given their persistence, bioaccumulative potential, and widespread emissions, PFASs represent a major public health challenge. The increasing volume of research documenting their toxicity has prompted regulatory bodies worldwide to implement stricter legal frameworks [ 1 , 5 ]. Understanding the physicochemical properties and primary exposure routes is essential for preventing further contamination and developing effective remediation strategies. This review aims to synthesize current knowledge on PFAS exposure sources and evaluate the efficacy of existing regulations based on observed pollution trends.

The pervasive environmental contamination by per- and polyfluoroalkyl substances (PFASs) presents an unprecedented public health crisis driven by the extreme thermodynamic stability of the carbon–fluorine (C–F) bond, which renders these compounds highly recalcitrant to conventional remediation. Empirical evidence demonstrates that toxicological risk assessments must pivot from reductionist single-compound models toward holistic mixture and co-exposure paradigms. While observational associations regarding general population carcinogenicity exhibit statistical heterogeneity, the epidemiological and experimental evidence strongly suggests an association with systemic multi-organ toxicities—specifically intrahepatic steatosis, chronic kidney disease progression, and barrier-surface immunosuppression. However, given the reliance on observational study designs and the inherent variability in exposure biomarkers, these findings must be interpreted with methodological caution, strictly distinguishing between statistical correlation and biological causality. Furthermore, the industrial substitution of legacy long-chain congeners with short-chain alternatives has backfired; these homologues possess heightened hydrogeological mobility, facilitate long-range transport, and resist standard granular activated carbon adsorption. This management crisis is severely compounded by international regulatory fragmentation, where the sharp divergence between stringent US EPA maximum contaminant levels (4 ng/L for PFOA/PFOS) and broader European or Asian thresholds compromises unified risk mitigation across globalized supply chains. To mitigate this transgenerational toxicological pressure, strategic interventions must prioritize class-based legislative frameworks aligning with the 2021 OECD expanded criteria, thereby eliminating the cycle of regrettable commercial substitutions. Concurrently, environmental engineering priorities must shift from purely segregative phase-transfer technologies toward advanced destructive modalities capable of complete C–F bond mineralization, such as electrochemical oxidation, non-thermal plasma, and supercritical water oxidation. Finally, future longitudinal cohorts must systematically integrate sex-specific metabolic dimorphisms and fetal epigenetic programming mechanisms, supported by targeted, high-resolution biomonitoring of critical occupational and pediatric populations.