Srinithi Mayilswamiab,
Nirav P. Ravalc,
Rinki Tomard,
Shailja Sharma
d,
Sarva Mangala Praveena
e,
Navish Katariaf,
Rangabhashiyam Selvasembiancg,
Saravanan Ramiah Shanmugam
h,
Ravinder Nathi,
Arindam Malakar
j,
Sudeshna Duttak and
Santanu Mukherjee
*il
aCentre for Environmental Risk Assessment and Remediation, University of South Australia, Building X, Mawson Lakes, SA 5095, Australia
bCooperative Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE), University of Newcastle, ATC Building, Callaghan, NSW 2308, Australia
cDepartment of Environmental Science and Engineering, School of Engineering and Sciences, SRM University-AP, Andhra Pradesh 522 240, India
dSchool of Biological & Environmental Sciences, Shoolini University of Biotechnology and Management Sciences, Solan 173229, India
eDepartment of Environmental and Occupational Health, Faculty of Medicine and Health Sciences, Universiti Putra Malaysia, UPM Serdang, Selangor Darul Ehsan 43400, Malaysia
fDepartment of Environmental Sciences, J. C. Bose University of Science and Technology, YMCA, Faridabad, 121006, Haryana, India
gCentre for Interdisciplinary Research, SRM University-AP, Amaravati, Andhra Pradesh 522240, India
hDepartment of Biosystems Engineering, Auburn University, Auburn, AL 36849, USA
iSchool of Agriculture Sciences, Shoolini University of Biotechnology and Management Sciences, Bajhol, PO Sultanpur, Distt., Solan, Himachal Pradesh 173229, India. E-mail: santanu@shooliniuniversity.com
jNebraska Water Center Part of Daugherty Water for Food Institute, School of Natural Resources, University of Nebraska-Lincoln, Nebraska, USA
kDepartment of Family and Community Medicine, Creighton University, CHI Health Creighton University Medical Center – University Campus, 2412 Cuming Street, Suite 102, Omaha, NE 68131, USA
lDepartment of Environmental and Civil Engineering, Faculty of Engineering, Toyama Prefectural University, 5180 Kurokawa, Imizu, Toyama 939-0398, Japan
First published on 24th September 2025
The widespread incorporation of per- and polyfluoroalkyl substances (PFAS) in various daily-use items has garnered considerable attention regarding environmental and health hazards in the last decade. Among different categories of PFAS, a paradigm shift has occurred towards short-chain PFAS alternatives like GenX, ADONA, and F53B, driven by environmental considerations and regulatory changes. Exposure to PFAS can happen through consuming contaminated food and drink, inhaling contaminated dust, or skin contact with PFAS-containing objects. Furthermore, occupational exposure might result from manufacturing and firefighting operations employing fluorinated compounds. In humans and monkeys, perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) exhibit an increased affinity for plasma proteins. However, the exact extent of this affinity is still a matter of research. The buildup of PFOS in the liver might cause injury or dysfunction by interfering with its regular operation. Compared to other human tissues, the liver has been shown to accumulate higher amounts of PFOS. Although there is an absence of epidemiological studies on PFOS, a possible connection between the health disorder and elevated cholesterol levels has been established by many researchers. Considering the transition as a future environmental burden, this review aims to bring together ongoing research compilations on short-chain PFAS, delving into their persistence, prevalence, and bioaccumulative toxicity in aquatic environments and focusing on critical areas of research gaps. An extensive literature analysis assessed the relative abundance of short-chain compounds compared to their long-chain counterparts within aquatic ecosystems. US EPA has setup new guidelines specifically for drinking water for PFOA and PFOS compounds which is 4 ppt. Furthermore, this review highlights emerging regulatory measures being implemented worldwide to safeguard public health. These measures encompass a range of strategies, from the European Union's emphasis on banning certain manufacturing and production practices under the REACH regulations to establishing exposure limits and disposal protocols in the United States.
Environmental significanceThe widespread incorporation of per- and polyfluoroalkyl substances (PFAS) in various daily-use items has garnered considerable attention regarding environmental and health hazards in the last decade. Among different categories of PFAS, a paradigm shift has occurred towards short-chain PFAS alternatives like GenX, ADONA, and F53B, driven by environmental considerations and regulatory changes. Exposure to PFAS can happen through consuming contaminated food and drink, inhaling contaminated dust, or skin contact with PFAS-containing objects. The significant of the current review is distinct from previous works, focusing specifically on consolidating recent literature concerning short-chain PFAS and providing insights into the regulatory measures implemented and adopted globally to address the persistent environmental presence and human health risks posed by PFAS. |
PFAS molecules are interconnected carbon and fluorine atoms, forming strong carbon-fluoride bonds. Due to this bond's resilience, these chemicals do not readily degrade in nature. PFOS and PFOA have garnered significant attention as persistent organic pollutants (POPs) because of their unnoticed existence in the environmental ecosystem, including human serum and tissues.2 They are classified as long-chain PFAS, distinguished by their eight-carbon backbone with sulfonate and carboxylate functional groups. Their ability to repeal water (hydrophobicity) and oil (oleophobic), combined with a variety of other chemical attributes, renders them valuable in a myriad of consumer goods.3,4
The adoption of shorter-chain PFAS over long-chain molecules in the early 2000s marked a shift towards using compounds with carbon backbones containing fewer than seven carbons in industrial and environmental applications.5 Noteworthy among these shorter-chain substitutes are GenX (hexafluoropropylene oxide dimer acid (HFPO-DA)), ADONA (4,8-dioxa-3H-perfluorononanoate), and F53B (chlorinated polyfluoroalkyl ether sulfonate),which have gained widespread usage.6 GenXis employed in various industrial processes.7 ADONA is applied as a replacement for PFOA in synthesizing fluoropolymers,1,8 whereas F53B serves as a replacement for PFOS and functions as a mist suppressant in electroplating processes.9,10 Despite extensive research on the toxicity and health impacts of PFAS, gaps remain in understanding the unique challenges posed by short-chain variants.11–13
While existing literature, including comprehensive reviews and empirical studies, has explored the effects of PFAS primarily in rodents, wildlife, fish, and through human autopsy cases, there is a critical need for deeper analysis into the impacts on other animal species and studies on human health.14–16 Furthermore, the effectiveness of global regulatory responses to PFAS, key emerging contaminants, exhibits significant variability across different jurisdictions. This inconsistency complicates efforts to mitigate the environmental and health impacts of PFAS and hampers the ability to conduct comprehensive, comparative analyses of policy efficacy.
Divergent regulatory landscapes, ranging from stringent prohibitions in some countries to more lenient guidelines in others, pose a challenge for multinational enforcement and global environmental protection strategies. Furthermore, the lack of uniform standards impedes the development of international agreements that could facilitate more effective management of PFAS pollution. The primary objective of the current review is distinct from previous works, focusing specifically on consolidating recent literature concerning short-chain PFAS and providing insights into the regulatory measures implemented and adopted globally, which is the novelty aspect of the current review. The specific objectives of this review are: (i) providing a comprehensive survey to compare the persistence and prevalence of short-chain molecules with their long-chain counterparts in aquatic environments, (ii) identifying knowledge gaps by conducting an in-depth literature review on the bioaccumulation of ADONA and GenX in various aquatic organisms and assessing potential eco-toxicological implications, and (iii) evaluating the human health risks connected with exposure to short-chain PFAS. Lastly, this review seeks to evaluate how these disparate regulatory frameworks influence PFAS management and control, aiming to identify best practices and recommend approaches for regulatory harmonization that could enhance global efforts to address the persistent environmental presence and human health risks posed by PFAS.
HFPO-DA (Gen-X) has been identified in surface waters in Germany and the Netherlands, particularly downriver from fluorochemical production units, reaching concentrations of ∼800 ng L−1.24,25 It was found during a comprehensive survey of surface waters across China, European countries, Korea, and the USA, showing 0.18 to 144 ng L−1 concentration levels.26–29 ADONA, identified in the Rhine River (Europe) water samples with a 75% detection frequency, generally exhibited low concentrations ranging from less than 0.01 to 1.5 ng L−1.30,31 These compounds' environmental presence is concerning, primarily due to their high mobility in soil and water systems.27 Furthermore, certain research findings indicate that their ultimate degradation byproducts exhibit persistence.32
Additionally, the data from 15 countries covering the occurrence of short-chain PFAS in various drinking water matrices (tap, bottled, and groundwater) was collected and tabulated. Table 1 includes 13 short-chain PFAS, with PFBA, PFPeA, PFHxA, PFBS, and PFHxS being the most frequently reported across global studies. The Netherlands reported the highest diversity with 13 short-chain PFAS detected in treated waters, including ultra-short-chain compounds like trifluoroacetic acid (TFA) and 6:
2 diPAP, with concentration ranges reaching up to 520.9 ng L−1 for TFA. China demonstrated the highest concentrations overall, with PFBA and PFHxA exceeding 9000 ng L−1 and 8000 ng L−1 respectively in drinking water sources near a former fluorochemical facility. Singapore, South Korea, and Norway reported moderate levels (typically <10 ng L−1), suggesting a relatively lower burden or effective regulatory mitigation. The US and Canada revealed a widespread but moderate-level presence of multiple SC-PFAS in bottled and municipal waters.
Country | City | No: of short chain PFAS | Name of short chain PFAS | Drinking water | Ground water (ng L−1) | References | ||
---|---|---|---|---|---|---|---|---|
Bottled water (ng L−1) | Drinking water sources (ng L−1) (households, municipalities) | |||||||
a BDL: below detection limit. | ||||||||
Singapore | Multiple locations (East, NortheastWest and Central Singapore) | 5 | • PFPeA | • 0.08–0.88 | • 0.11–3.51 | — | 124 | |
• PFHxA | • 0.06–0.94 | • 0.21–4.63 | ||||||
• PFHpA | • 0.07–0.17 | • 0.07–1.79 | ||||||
• PFBS | • 0.06–0.83 | • 0.07–1.99 | ||||||
• PFHxS | • 0.08–0.14 | • 0.06–0.97 | ||||||
China | Hubei province | 5 | • PFBS | — | • BDL – 31.0 | • 6.4–9238.5 | 125 | |
• PFHxS | • BDL – 160.1 | • BDL – 5999.2 | ||||||
• PFPeA | • BDL – 0.6 | • BDL – 1259.8 | ||||||
• PFBA | • BDL – 2.4 | • 13.0–8639.5 | ||||||
• PFHxA | • BDL – 3.4 | • 0.1–3351.4 | ||||||
South Korea | Multiple location | 4 | • PFPeA | • 12.7 | — | — | 126 | |
• PFHxA | • 12.9 | |||||||
• PFHxS | • 5.92 | |||||||
• PFBS | • 5.68 | |||||||
Combine | Mainly Canada (n = 95), USA (n = 22), France (n = 9) | 3 | • PFBA | — | • 0.113–104.6 | — | 127 | |
• PFPeA | • 0.084–87.61 | |||||||
• PFHxA | • 0.043–84.41 | |||||||
• Gen-X | • 0.198–5.28 | |||||||
Belgium | Flanders | 4 | • PFBA | — | • BDL – 12 | • 21 | 128 | |
• PFBS | — | • BDL – 7.3 | ||||||
• PFPrS | — | • BDL – 0.6 | ||||||
• TFMS | BDL – 15 | • BDL – 5.6 | • 9.4 | |||||
Norway | Various municipality water suppliers | 7 | • PFBA | — | • 0.019 | — | 129 | |
• PFPA | • 0.11 | |||||||
• PFPrS | • 0.052 | |||||||
• PFBS | • 0.077 | |||||||
• PFPS | • 0.0041 | |||||||
• 6![]() ![]() |
— | |||||||
• PFHxA | • 0.083 | |||||||
Italy | Vicenza | 3 | • PFBA | — | — | • 61 | 130 | |
• PFPeA | • 38 | |||||||
• PFHxA | • 38 | |||||||
Netherlands | Different treatment plants | 13 | • TFA | — | • 88.44–482.95 | • 87.7–520.9 | 131 | |
• PFPrA | • 0.12–28.39 | • 0.12–11.13 | ||||||
• PFBA | • 0.3–1.17 | • 0.03–3.69 | ||||||
• PFPeA | • 0.03–0.3 | • 0.03–0.54 | ||||||
• PFHxA | • 0.03–0.35 | • 0.19–0.54 | ||||||
• PFBS | • 0.02–0.12 | • 0.02–0.1 | ||||||
• PFPeS | • 0.03–0.09 | • 0.03–0.17 | ||||||
• 4![]() ![]() |
— | — | ||||||
• 4![]() ![]() |
— | — | ||||||
• PF40PeA | — | — | ||||||
• 6.2 diPAP | — | — | ||||||
• ADONA | — | — | ||||||
Czech Republic | Various Household around Czech Republic | 4 | • PFHxA | — | • 0.048–97.7 | — | 132 | |
• PFBS | • 0.093–3.10 | |||||||
• PFPeS | • 0.049–0.699 | |||||||
• PFHxS | • 0.026–0.772 | |||||||
US | Baltimore metropolitan area | 7 | • PFPrA | • 0.45–6.52 | — | — | 133 | |
• PFBA | • 0.51–1.40 | |||||||
• PFBS | • 0.19–1.44 | |||||||
• PFPeA | • 2.84–3.27 | |||||||
• PFPeS | • 0.21–0.23 | |||||||
• PFHxA | • 0.41–1.98 | |||||||
• PFHxS | • 0.31–0.64 | |||||||
Turkey | Multiple locations | 4 | • PFBA | • 0.26–0.46 | • 0.27–1.93 | — | 134 | |
• PFPeA | • 0.08–0.14 | • 0.08–1.23 | ||||||
• PFHxA | • 0.08–0.21 | • 0.08–2.90 | ||||||
• PFBS | • 0.19–0.21 | • 0.11–0.85 | ||||||
India | Different points of Ganges | — | • PFHxA | — | — | • 0.8–4.9 | 135 | |
• PFBS | — | |||||||
• PFBA | • BDL – 9.2 | |||||||
• PFHpA | • 0.5–3.5 | |||||||
Brazil | Porto Alegre metropolitan area & other locations | 4 | • PFBS | • 3.1–3.6 | • 0.48–16 | — | 136 | |
• PFHxA | — | • BDL – 15.9 | ||||||
• PFHxS | — | — | ||||||
• PFHpA | • 5.7–6.8 | • 5.2–36 | ||||||
Spain | Barcelona metropolitan area& other locations | 4 | • PFBS | — | • 2.8–24 | — | 136 | |
• PFHxA | • 4.8–11.8 | • 14–58 | ||||||
• PFHxS | — | — | ||||||
• PFHpA | • 5.7–17 | • 4.1–42 | ||||||
France | Toulouse, Montpelier, Nimes, Avignon, Valence, Grenoble, Lyon and Perpignan | 4 | • PFBS | • 1.3–6.7 | • 2–15 | — | 136 | |
• PFHxA | — | • 5.8–6.8 | ||||||
• PFHxS | • 6.7 | — | ||||||
• PFHpA | • 4.5–25 | • 13–33 |
Countries such as India, Brazil, and Spain showed notable occurrence of PFBS and PFHxA in both groundwater and household drinking water sources, indicating ongoing exposure risks in developing and middle-income regions (SI Fig. S3).
However, data regarding the adverse impacts of PFAS on other animal species and humans remains limited (Fig. 2). Studies have advocated that the long half-lives of PFAS in humans may be due to their strong binding to plasma proteins.34,35 Both PFOA and PFOS demonstrate a heightened attraction to plasma proteins in monkeys and humans, although the precise degree of binding remains incompletely characterized. PFOS tends to accumulate more prominently in the liver and serum, carrying significant implications.36 Accumulation of PFOS in the liver can disrupt its normal functioning and potentially lead to liver damage or dysfunction.37 Higher levels of PFOS accumulation in hepatic tissues maybe attributed to enterohepatic recirculation, wherein PFOS is excreted in bile and subsequently reabsorbed from the gut.38 In the bloodstream, PFOS in the serum can circulate throughout the body, potentially affecting various organs and systems. During the 1990s, studies conducted in the United States revealed that serum samples from pooled blood banks had average PFOS concentrations ranging from 28 ng g−1 to 44 ng g−1.39 Similarly, research conducted in Europe reported mean serum PFOS concentrations of 17 ng g−1 in Belgium, 53 ng g−1 in the Netherlands, and 37 ng g−1 in Germany, all based on pooled blood bank samples.40–42 Among 23 human donors with paired samples, it was found that the mean PFOS concentrations were 20.8 ng g−1 in the liver and 1.32 μg mL−1 in the serum.39
Several reviews have examined the prevalence of PFOA and its involvement in a large population's health hazards, such as immunomodulation.43,44 The earlier studies consistently observed moderate increases in cholesterol and uric acid levels with PFOA. However, results regarding long-term disorders with an inflammatory component, such as diabetes and stroke, were inconclusive. The same authors also reported reproductive and developmental disorders in humans.43 The study concluded that low birth weight resulting from PFOA exposure was typical and did not carry significant clinical implications. Granum et al. (2013)45 also reported alterations in serum immunoglobulin levels, and male PFOA workers showed an association with elevated monocyte counts in residents exposed to PFOA-contaminated water, as Brieger et al. (2011)46 reported. While epidemiological data on PFOS is limited, existing information suggests a potential link to health disorders associated with increased cholesterol levels.
PFASs | Population samples | Exposure route measurements | Countries | Impacts on human life | References |
---|---|---|---|---|---|
PFOA | Adults (54% women) (61% men) | Plasma | U.S state | • PFAS promoted liver disease and hepatocellular apoptosis through dysregulation of caspase −3 enzyme | 137 |
PFOS | |||||
PFHxS | West (Virginia) | • Exposure to PFHxS and PFOA increased the activation of C3a peptide in men and reduced it in women | |||
PFNA | |||||
PFOA | Children (8 years) | Venous blood | U.S state (Ohio) | • PFAS has been shown to be associated with biological pathways such as proline, aspartate butanoate, and asparagine | 138 |
PFOS | |||||
PFNA | |||||
PFHxS | |||||
PFOA | Children (8–14 years) | Plasma | America (Los Angeles) | • Significant dysregulation of several amino acids and lipids included tyrosine, proline, arginine and linoleic acid de novo lipogenesis observed with exposure of PFAS in children | 139 |
PFOS | |||||
PFHxS | |||||
PFOA | Children (7–19 years) | Plasma | United States (Atlanta) | • Elevated concentration of PFAS caused liver fibrosis and lobular inflammation | 140 |
PFOS | |||||
PFHxS | • Hepatocellular ballooning was found in 40% of children | ||||
PFOS | Pregnant women | Cord blood | China | • Exposure to PFAS has been shown to affect fetal growth such as body weight, head circumference, and body length at birth | 141 |
PFOA | • Impacts on estragon homeostasis also showed | ||||
PFNA | Children (0, 3 and 7 years) | Plasma | Norway | • Exposure to PDUnDAstudied to havea higher incidence of Atopic eczema or atopic dermatitis in girls children | 142 |
PFHxS | |||||
PFOA | • Throat infection, pseudo-croup bronchitis were observed in children at ages of 0-, 3- and 6-years kids | ||||
PFOS | • The majority of girls child found to be infected | ||||
PFUnDA | |||||
PFOS | Children (7–8 years) | Plasma | United States (Boston) | • BMD affected by exposure to PFOS and PFOA substances | 143 |
PFOA | |||||
PFOS, PFOA | Pregnant women | Blood sample | Northern Norway | • The deregulation of thyroid homeostasis has been studied | 144 |
PFNA, PFHxS | |||||
PFUnDA | • Fatal health have been affected by higher exposure to PFAS | ||||
PFDA | |||||
PFHxS, PFOS | Children (5, 7 and 13 years) | Blood sample | Faroe Islands | • Children without given MMR vaccination at the age of 5 years, had the risk of allergic diseases and asthma | 145 |
PFOA, PFNA | |||||
PFDA | |||||
PFHxS | — | Human cell line (H295R, MCR-7and LNCaP) | Germany | • Contact with PFOS and PFOA has been shown to induce estragon receptor activities and enhanced Astron secretion in H295R cells | 146 |
PFBA | |||||
PFHxA | • PFOA increased secretion of progesterone at higher concentrations (100 μM) however, <10 μM concentration did not showed any harmful impacts | ||||
PFBS | |||||
PFNA | Adults (18–38 years) | Blood and urine | Fernald | • Dysfunction of kidneys and thyroid observed in that study | 147 |
PFHxS | |||||
PFDeA | |||||
PFOA | |||||
PFOS | |||||
PFOA | — | Human hepatocytes cells | Germany | • Liver cells dysfunctionalities were found at high concentration (100 μM) of PFOA exposure to hepatocytes cells | 148 |
PFOA | Adults | Semen sample | China | • Spontaneous acrosome reaction did not affect with contact of PFOA at 0.25 to 2.5 μg ml−1 concentrations | 149 |
PFDA, PFOS | Women | Plasma | China | • In Chinese women, PFAS exposure has been linked to higher incidence of infertility | 150 |
PFOA, PFUA | |||||
PFDoA | |||||
PFNA | |||||
PFBS | |||||
PFTA | Men and women | Blood sample | Italy | • PFAS contaminated drinking water, caused cerebrovascular diseases, diabetes, Alzheimer's disease and myocardial infarction | 151 |
PFOS | |||||
PFOA | • Risk of breast cancer and kidney infection observed | ||||
PFDA | |||||
PFBS | |||||
PFOS, PFOA | Adults (18–25 years) | Blood lipids | U.S state | • A longitudinal study found that PFAS increased the risks of diabetes | 152 |
PFHxSEtFOSAA | • Hypertriglyceridemia and hypercholesterolemia risks were also associated to exposure of PFAS | ||||
MeFOSAA | |||||
PFOA | Children (2–5 years) | Blood sample | America | • Children had thyroid dysfunctions (elevated FT4 level and reduction in THS) to exposure of PFAS through breastfeeding at early life stage | 153 |
PFOS | |||||
PFHxS | |||||
PFNA | |||||
PFDN | |||||
PFOA | Mother and child pair | DNA methylome | Europe | • That study found that PFAS exposure may be caused significant respiratory diseases in children | 154 |
PFOS | Pregnant women (mother–child pairs) | Blood sample | America | • Study analysed symptoms of cerebral palsy (neurological disorder) in infants to high exposure level of PFAS through their mothers | 155 |
PFOA | |||||
PFNA | Children (8–12 years) | Blood sample | U.S state (Ohio) | • Exposure to PFNA has been showed negative association with bone health | 11 |
PFOA | |||||
PFOS | • Low density lipoprotein cholesterol and increased of systolic blood pressure was observed | ||||
PFHxS | |||||
PFOS | Pregnant women | Maternal fasting blood sample | U.S State (Colorado) | • Lower weight was observed in females prenatal with the association of PFHxS and PFOS | 156 |
PFHxS | • Exposure to MePFOSAA, in early life stage, affected adiposity | ||||
MeFOSAA | |||||
PFOS | Pregnant women | Pregnant women | Denmark | • Exposure to PFOA, PDFA, PFHxS and PFHpS were linked with infant girls while PFHpS and PFHxS with infant boys, but no significant impact was observed | 157 |
PDFA | |||||
PFHxS | |||||
PFHpS | |||||
PFHxS | Women | Blood sample | Canada | • Exposure to PFOS and PFHxS reduced fecundability (ability to conceive a pregnancy) | 158 |
PFOS | |||||
PFOS | 172 mother–child pairs | Maternal cord serum | Faroe Islands | • PFAS has been shown negative association with head circumference, body weight and height | 159 |
PFOA | • PFOA and PFOS were responsible to elevated THs concentration, but did not show any impact on birth weight | ||||
PFOA, PFHxS | Men and women (20–40 years) | Blood sample | Korea (Seoul) | • Study analyzed positive association between PFAS and low density lipoprotein cholesterol, total cholesterol and triglycerides | 160 |
PFHpA | |||||
PFHxA, PFBS | |||||
,PFOS | • Higher concentration of PFOA and PFOS was observed in diabetic participants as compared to non-diabetic | ||||
PFUnDA | |||||
PFDS, PFDoDA | |||||
PFTrDA | |||||
PFOA | Children (boys and girls, 10–16 years) | Blood sample | Norway | • PFHPA caused asthma in girls at age of 10 years | 161 |
PFHpA | |||||
PFDA | |||||
PFUnDA | • Atopic dermatitis observed in girls to exposure of PFUnDA, PFDA and PFOA in boys | ||||
PFHxS | |||||
PFHpS | • In a cross- sectional study from Norway, it was observed that girls had higher risk of allergies than boys | ||||
PFOS | |||||
PFOS | Pregnant women | Maternal plasma | Canada | • In Canada, a cohort research discovered that PFAS had no significant relationship with gestational weight growth | 162 |
PFOA | |||||
PFHxS | |||||
PFOS, PFOA | Pregnant women (early stage of pregnancy) | Blood sample | China (Shanghai) | • Exposure to PFAS did not show any positive association with PE (preeclampsia) and GH (gestational hypertension) at birth time | 163 |
PFHUnDA | |||||
PFHxS, PFDA | |||||
PFBS, PFHpA | |||||
PFOSA, PFDoA | |||||
PFOS | Men and women (>20 years) | Blood sample | United States | • The risk of diabetes in men was studied due to exposure of PFOA | 164 |
PFHxS | • However, PFAS has been shown association with total cholesterol observed in adult candidates | ||||
PFOA | |||||
PFOA | Male workers (PFs factory) | Blood sample | Italy | • Toxic impacts such as liver cirrhosis, liver cancer, diabetes and malignant neoplasm were observed in PFs factory's workers due to high exposure of PFOA | 165 |
PFOS and PFHxS | Mother child pair | Blood sample and questionnaires | Japan (Hokkaido) | • Longitudinal study found immunotoxicity due to exposure of PFAS in offspring | 166 |
• PFHxS caused a high risk of infections in girls children | |||||
• Otitis media, respiratory syncytial virus and pneumonia diseases were observed due to exposure of PFAS in children at age of 4 years | |||||
PFOA and PFOS | Male adults (10–21 years) | OUS (quantitative ultrasounds) | Italy | • Elevated PFAS exposure may raise the incidence of osteoporosis (bone weakening) in men between the ages of 18 and 20 | 167 |
PFOA | Pregnant women and mother–newborns pairs | Cord blood | China (Wuhan) | • PFOA and PFOS were positively corelated with 11- deoxy cortisol, cortisol and progesterone in newborns | 168 |
PFDoDA, PFUnDA, PFOS, PFBS | |||||
PFDA | |||||
PFOS, PFOA | Adults (>20 years) Obese and non- obese | Blood sample | United States | • PFAS had impact on dysregulation of lipid biomarkers, including LDA and TC in obese participants | 169 |
PFDA | • However, PFNA, PFHxS, PFOA were found to be positively correlated with caline aminotransferase in obese participants | ||||
PFHxS | |||||
PFHxS | Men, women and children (12–80 years) | Venous blood | United States | • Exposure to PFAS may lead to elevated TSH in male and reduced it in females during adolescence | 170 |
PFOS, PFOA | • PFAS may be responsible to increased TT3, FT3 and FT4 in females | ||||
PFOA | Men and women (70, 75 and 80 years) | Blood sample | Sweden | • A longitudinal study revealed that PFDA, PFOA, PFOS, PFHpA were positively linked with liver enzymes like ALT and ALP | 171 |
PFDA, PFDnDA | |||||
PFDoDA | |||||
PFHpA |
PFOA exposure has been associated with thyroid diseases, and epidemiological research suggests a possible association with human cancers.48 Young men experienced a decline in semen quality after coming in contact with PFOS, PFHxS, and PFOA.49 Furthermore, these three chemicals have been related to early onset menopause in females, increased impulsivity, and delayed puberty in children.50,51 The presence of PFHxA and PFBA in human autopsy tissues has revealed distinct patterns, with PFHxA predominantly detected in the brain and liver, while PFBA is frequently noticed and found at higher concentrations in the kidney and lung.52
Sub-chronic hepatotoxicity of Cl-PFAES in mice was observed, with fatty liver and indications of cell apoptosis and proliferation in groups exposed to doses exceeding 0.2 mg per kg per day.55 Mice exposed to GenX showed a higher occurrence of placental abnormalities. In contrast, affected rats showed higher expression of peroxisome proliferator-activated receptor (PPAR)-regulated genes in both livers, resembling the impacts of PFOA noted previously. GenX exhibited developmental toxicity in rats, leading to higher rates of neonatal deaths and lower birth weights in those exposed from gestational day 8 to postnatal day 2, with doses ranging from 1 to 125 mg per kg per day.56,57
Studies conducted using rat neuron cultures have elucidated that the impact of PFCs is contingent upon their molecular arrangement, characterized by a carbon chain ranging from 4 to 16 atoms, enclosed by fluorine atoms. These compounds often contain a charged functional group, such as carboxylate, sulfonate salt, or acid, at one end.59 Consequently, they have been observed to markedly elevate ROS formation within cells, potentially by activating pathways such as PPARα or nuclear factor erythroid 2-related factor 2 (Nrf2). Although these effects may not directly precipitate cell death, they induce oxidative stress, damage DNA, and cause various physiological alterations. While long-chain PFAS like PFOA and PFOS have been well-documented for their cytotoxic effects,60 few studies suggest that short-chain PFAS such as PFBS and PFHxA also induce oxidative stress, although at typically higher concentrations and with lower bioaccumulative potential.61,62 This suggests reduced, but not negligible, toxicity for shorter-chain analogs.
The concentrations of PFAS associated with adverse health impacts in humans range from 2 to 20 ng mL−1, as reported by the NASEM in 2022.63 However, individuals with occupational exposure to PFAS may have higher serum levels, with typical levels around 300 ng mL−1 for PFOS and approximately 2000 ng mL−1 for PFOA.64
Experiments on HepG2 cells have shown that PFOA can induce genotoxic effects. Research by ref. 66 and 67 suggested that among various PFCs, only those with eight or nine carbon atoms in their structure were capable of producing ROS or causing DNA damage in HepG2 cells. However, the observed impact was mild, and a clear dose–response relationship was not established.68,69 PFCs also cause oxidative damage associated with nuclear receptor proteins that regulate gene expression.70
The carcinogenic potential of PFOA has been shown in human MCF-7 breast cancer cells due to its estrogen-like properties, indicating the endocrine-disrupting capabilities of PFCs.72 Additionally, PFOA and perfluoro-n-decanoic acid (PFDA) have been linked to an increased risk of breast cancer in Chinese women. This association is attributed to the disruption of hormone balance caused by the combined xenoestrogenic and xenoandrogenic activities of serum POPs, thereby increasing susceptibility to breast cancer.73
A study investigated the prevalence of cancer among workers of a perfluorooctane sulfonyl fluoride (POSF) manufacturing facility, finding that employees with higher exposure had a higher incidence of deaths due to bladder cancer.74 However, there is still debate regarding a clear link between PFOA exposure and human disease incidence.75 Studies on rodents have shown that PFOS and PFOA act as peroxisomal proliferators. Notably, peroxisome proliferators' carcinogenic potential does not appear to impact humans.71
The potential mechanisms of exposure to PFAS and cancer development can be broadly categorized into four significant aspects. Firstly, PFAS compounds can influence hormone receptors and disrupt the delicate balance of the endocrine system, leading to alterations in hormone receptors and potential hormonal imbalances.76 Secondly, PFAS have been found to activate a specific receptor known as PPARα. This activation can induce oxidative stress within the body, a state with excess harmful ROS. Oxidative stress is known to be detrimental to cellular health and can contribute to the development of cancer.77 The third mechanism involves PFAS-induced epigenetic alterations, encompassing changes in DNA methylation patterns. It also modifies gene expression-regulating proteins, mostly histones. PFAS exposure-induced epigenetic changes may contribute to tumorigenesis.78 Lastly, PFAS are associated with reproductive toxicity, potentially increasing susceptibility to carcinogens and impacting breast cancer risk.
The mechanism behind PFOS-induced neonatal death is currently unknown. However, PFOS targets organ systems that develop during the later stages of pregnancy, which aligns with previous research and teratological outcomes.83 Additionally, PFOS-mediated organ failure contradicts postnatal survival, suggesting that lung development and pulmonary function may be a significant point of impact.84 Instances of neonatal mortality linked to PFOS exposure share similarities to the effects of nitrofen, an herbicide known for interfering with fetal lung growth, leading to compromised cardiopulmonary function and increased neonatal rat mortality.85 The research studies indicate observable changes in lung structure and size in PFOS-exposed newborns, resulting in hindered development of the lungs during the perinatal period.
As presented in Fig. 3(a), PFAS exposure has been associated with multiple adverse effects on the male reproductive system. Research has shown that PFAS compounds can interfere with hormone regulation, leading to disruptions in testosterone levels and sperm quality. An imbalance in hormone levels can lead to problems like lower sperm count, reduced sperm movement, and changes in sperm shape. Additionally, PFAS exposure has been related to testicular damage and dysfunction, including testicular atrophy and impaired spermatogenesis. These effects can ultimately lead to fertility problems and reproductive disorders in males. PFAS exposure can also impact the female reproductive system in several ways, as presented in Fig. 3(b). Similar to males, PFAS can disrupt hormone regulation in females, affecting estrogen and progesterone levels. This hormonal imbalance can lead to menstrual irregularities, reduced fertility, and difficulties in conceiving. Compared to long-chain PFAS, short-chain alternatives like show weaker binding to hormone receptors and reduced transplacental transfer; however, some studies still report endocrine disruption and fetal development concerns at elevated doses.86 Furthermore, prenatal exposure to PFAS has been correlated with reduced birth weight and elevated preeclampsia risk. Studies have also suggested a potential connection between PFAS exposure and an increased incidence of gynecological conditions like endometriosis and ovarian cysts.87
Exposure to PFOA has been found to disrupt hormonal balance in rodents, resulting in Leydig cell hyperplasia and the formation of Leydig cell adenomas.97 Studies on adult rats treated with greater than 5 mg perfluorododecanoic acid (PFDA) per kg body weight daily for two weeks have observed testicular damage, along with changes in gene expression, particularly the suppression of genes responsible for cholesterol transport and steroidogenesis, as well as a decrease in serum testosterone levels.98 These alterations are concerning, as Leydig cell hyperplasia is commonly observed in impotent men with lower testosterone levels compared to normal individuals.99 Improper testicular function is associated with testicular dysgenesis syndrome (TDS).100 It is believed that TDS is caused by exposure to endocrine disruptors during fetal development, which can affect testis formation and lead to impaired testicular function in adulthood, including reduced semen quality.
Research conducted on rats during crucial developmental periods has revealed testis dysgenesis marked by Leydig cell hyperplasia and Leydig cell aggregation in the testis center. This condition leads to decreased testosterone levels and reduced fertility in adulthood.101,102 The impaired function of Leydig cells results in the reduced expression of genes responsible for steroidogenesis.103
Numerous investigations have delved into PFOS's capacity to bind with serum proteins, studying how it displaces steroid hormones from specific binding proteins in birds' and fish's serum. PFOS is constrained to displace estrogen or testosterone from carp serum steroid-binding proteins. However, it disrupts cortisone in avian sera at relatively low PFOS concentrations. Additionally, the disruption of corticosterone increases with the chain's length, with sulfonic acids being more effective than carboxylic acids.93 The earlier research, which was focused on assessing the endocrine-disrupting potential of pollutants, mainly through non-receptor-mediated pathways, confirmed the interaction of contaminants with serum steroid-binding proteins as a potential mechanism. Previous research indicates that environmental pollutants like bisphenol A (BPA) and nonylphenol have limited efficacy in displacing human sex hormone-binding globulin (SHBG) ligands from SHBG to E2. Moreover, these contaminants have been found to increase the proportion of SHBG-unbound estradiol at 10 to 100 mM concentrations.108,109
Earlier investigations have indicated that humans exhibit an extended half-life for the serum removal of PFOS, PFHS, and PFOA.110 Species-specific differences in pharmacokinetics may be attributed to saturable renal resorption mechanisms. The average duration of serum elimination was around 5.4 years and 4.8 years for PFOS, 8.5 years and 7.3 years for PFHS, and 3.8 years and 3.5 years for PFOA.110 The extended half-life in humans for removing these substances might be due to differences in how they are excreted through bile and absorbed in the gut, possibly influenced by enterohepatic circulation.111 Table 3 presents various human health effects of PFAS with detailed experimental insights.
PFAS types (purity %) | Models used | PFAS doses (mg per kg per day) | PFAS exposure time (days) | Key conclusions | References |
---|---|---|---|---|---|
PFOS* (≥98%) | Sprague–Dawley rats | PFOS – 0, 2, 20, 50 and 100 | 28 | PFOS disrupts lipid balance and affects the endocrine system by reducing important hormones; however, it does not significantly impact the kidneys or cardiovascular system. PFOS primarily accumulates in the liver, with smaller amounts in the spleen and heart | 172 |
PFOS (98%) | Sprague–Dawley rats | PFOS – 5 and 20; PFOA – 5 and 20 | 28 | Abnormal behaviour, substantial weight loss, and an enlarged liver were noted in the high-dose PFOS-exposed rats, with PFOS accumulation highest in the liver, followed by the heart, kidney, whole blood, lung, testicles, spleen, and brain, in that order | 173 |
PFOA* (96%) | |||||
Potassium PFOS (87%) | C57BL/6 Mice | 6, 12 and 24 | 23 | Dietary PFOS led to a dose-dependent reduction in body weight, an increase in liver weight relative to the body, higher liver triglyceride levels, and elevated markers in the bloodstream, all pointing to liver toxicity and oxidative stress | 174 |
PFOS (>98%) | Sprague–Dawley rat | 3 | 7 | In PFOS-treated groups, liver weights were notably higher relative to body weights, with reduced serum thyroxine (TH) and no change in thyroid-stimulating hormone (TSH) levels. PFOS exposure caused lipid droplet formation, signifying TH disruption in both animals and humans, resulting in lower T4 and T3 levels in rats without TSH compensation | 175 |
PFOS (98%) | C57BL/6 (H-2b) mice | PFOS and PFOA – 2, 10, 40 | 10 | When comparing the impact of PFOS and PFOA on mice, both compounds, when administered at the same dose and duration (40 mg kg−1 for 10 days), led to similar effects. These effects included liver enlargement (hepatomegaly), reduced body weight, decreased thymus and spleen weights, fewer cells in the thymus and spleen, which also affected different cell subpopulations related to the immune system, and structural changes in the thymus | 176 |
PFOA (96%) | |||||
PFOS (98%) | CD-1 mice | 1, 5 and 10 | 21 | Mitochondria dysfunction and the elevation of oxidative stress could promote the development of steatohepatitis | 177 |
PFOS (98%) | C57BL/6J mice | 2.5, 5, and 10 | 30 | PFOS exposure induced hepatomegaly with dose-dependent increases in liver weight. Histopathology showed liver damage, edema, hepatocellular necrosis, and inflammation in PFOS-exposed mice | 178 |
PFOS | Sprague–Dawley rats | 1 and 10 | 14 | The findings of this study revealed that a dose of 10 mg kg−1 of PFOS in rats resulted in cardiac toxicity, characterized by heightened apoptosis and an upregulation of proinflammatory cytokines | 179 |
PFOA (>98%) | Kunming mice | 1, 2.5 and 5 | — | The current study found that maternal PFOA exposure had a minor influence on the testicular index of offspring mice. However, as the PFOA dosage increased the testes suffered varied degrees of damage | 28 |
PFOS | Sprague–Dawley rats | 0.5; 1.0; 3.0 and 6.0 | 28 | PFOS exposure disrupts the male reproductive axis by affecting the hypothalamus, pituitary gland, and testis. This disruption involves changes in noradrenaline concentration, gonadotropin-releasing hormone (GnRH) gene expression, and hormone secretion | 180 |
PFOS | Sprague–Dawley rats | 5 and 10 | 21 | Rats exposed to low dosages of PFOS during puberty may experience significant delays in the formation of Leydig cells, which is caused by the interruption of Leydig cell-specific gene expression. Furthermore, PFOS exposure is associated with lower seminal vesicle weights and reduced sperm counts | 181 |
PFOS | C57 mice | 0.5 and 10 | 35 | The current study showed that PFOS can reduce sperm production in mice. This effect is linked to a decrease in germ cell proliferation and an increase in apoptosis within the testis. These observations indicate that PFOS-induced testicular toxicity is influenced by estrogen receptors (ERs) | 182 |
PFOS | ICR mouse | 0.25 and 50 | 28 | PFOS can damage sertoli cells and weaken the blood-testis barrier (BTB), which can allow PFOS to pass through the testes and cause abnormalities in male reproduction | 183 |
PFOS (≥98%) | ICR mice | 0.5, 5 and 10 | 28 | The data indicated that PFOS led to a substantial reduction in sperm count and compromised the integrity of the blood-testis barrier (BTB) | 184 |
PFOA | Pregnant CD-1 mice | 1, 3, 5, 10, 20, and 40 | 17 | PFOA exhibits maternal and developmental toxicity in mice which causes early miscarriages, reduced survival after birth, delays in the growth and development of the whole body, and sex-specific changes in pubertal maturation, where male offspring show faster sexual maturation than female offspring | 185 |
PFOA (96%) | BALB/c mice | 0.31, 1.25, 5 and 20 | 28 | Male reproductive function may be adversely affected by PFOA exposure, as it interferes with testosterone levels and causes damage to the seminiferous tubules, resulting in decreased testosterone levels in the testes and an increase in spermatogonial apoptosis | 186 |
PFOS | Pregnant Sprague–Dawley rats | 5 and 20 | 19 | PFOS acts as an endocrine disruptor, lowering testosterone synthesis and altering fetal Leydig cells in rats, which affects the male reproductive system | 187 |
PFOS | Pregnant Sprague–Dawley rats | 1 and 5 | 20 | PFOS exposure reduced serum testosterone levels in a dose-dependent manner, potentially affecting Leydig cell development in offspring, and additionally, it decreased adrenal hormone aldosterone | 188 |
PFOS (>98%) | C57BL/6 mice | 1, 5, and 10 | 7 | Mice exposed to PFOS had a greater concentration of apoptotic cells compared to the control group. This elevated apoptotic cell count was linked to the generation of reactive oxygen species (ROS) by PFOS, which led to the dissipation of mitochondrial membrane potential and triggered apoptosis in splenocytes and thymocytes. Additionally, PFOS exposure resulted in increased activities of glutathione reductase, catalase, and superoxide dismutase | 189 |
PFOS (>98%) | Sprague–Dawley rats | 0.1 and 2.0 | 21 | The prenatal exposure to PFOS can disrupt the balance of antioxidant systems, leading to oxidative stress, and trigger caspase-dependent death pathways in the lungs of rat offspring | 190 |
PFOS | C57BL/6J mouse | 0.2, and 2.0 | 6 months | Mice exposed to PFOS may experience cognitive deterioration, this could be due to the activation of a particular pro-apoptotic pathway generated by endoplasmic reticulum stress in the cerebral cortex neurons | 191 |
PFOS | Sprague–Dawley rats | 1 and 5 | 18 | The gestational exposure to PFOS can lead to lung issues in offspring like bronchopulmonary dysplasia (BPD), potentially contributing to the rise in developmental lung diseases | 192 |
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Fig. 5 Short-chain PFAS contamination, pathways, impacts and regulatory framework in drinking water. |
The European Union (EU) has taken significant steps to address PFAS through various policy instruments. The EU Registration, Evaluation, Authorization, and Restriction of Chemicals (EU REACH) proposed restrictions on PFAS on February 7, 2023.113 These restrictions were targeted at the manufacturing, usage, and trade of these ECs. After that, PFAS substances have been banned unless their concentration is below 25 ppb for C9-C14 PFCAs and their salts or 260 ppb for C9-C14 PFCA-related substances.
Under REACH, PFAS substances are increasingly being classified as Substances of Very High Concern (SVHC), and restrictions have been proposed on their manufacture, use, and trade. This list has seen the addition of some short-chain groups of PFAS, which consist of substances like 2,3,3,3-tetrafluoro-2-(heptafluoropropoxy) propionic acid and their salts, acyl halides, PFBS, and PFHpA and their salts. In May 2020, Denmark prohibited the application of PFAS in food contact materials (FCMs). This law has been in effect and stipulates that paper and cardboard FCMs cannot be sold. Such targeted national bans have corresponded with measurable reductions in PFAS concentrations in sewage sludge, surface waters, and wildlife. However, it does allow for exceptions regarding PFAS use in FCMs if these products incorporate a functional barrier effectively preventing food contamination through PFAS.114
On February 20, 2025, the French Parliament passed pioneering legislation to phase out PFAS, targeting their widespread use and environmental persistence. The law bans PFAS in cosmetics, textiles, ski wax, and footwear by 2026, extending to all textiles by 2030, with exemptions for protective gear. It mandates PFAS monitoring in drinking water and introduces the “polluter pays” principle to hold polluting companies financially accountable. This legislation, following Denmark's example, positions France as a leader in PFAS regulation and may influence broader EU policy, advancing a unified framework for managing these hazardous substances across member states.115 Additionally, under the EU POPs Regulation, PFOS and PFOA compounds are strictly limited to trace amounts (0.001% by weight for PFOS and 0.0000025% by weight for PFOA), with exemptions allowed only for laboratory research or unintentional contamination.114
Within the EU, the principal legal framework overseeing water quality and access for human use is the Drinking Water Directive. This directive categorizes PFOS and related compounds as priority substances under water policy. A recent update in 2020 introduced new criteria, setting the ‘PFAS Total’ threshold at 0.5 μg L−1 and a maximum of 0.1 μg L−1 for the ‘Sum of PFAS’ in drinking water.116
In October 2021, the US-EPA unveiled its PFAS Strategic Roadmap. The title for this was EPA's Commitments to Action 2021–2024, and it details a comprehensive strategy comprising 31 specific initiatives falling under the EPA's regulatory purview.117 These initiatives are structured to be implemented over varying timeframes, encompassing discrete and ongoing projects. Key unresolved action points include:
On April 10, 2024, the U.S. EPA finalized drinking water standards for six PFAS under the National Primary Drinking Water Regulation (NPDWR). The rule sets legally enforceable Maximum Contaminant Levels (MCLs) at 4 ppt for PFOA and PFOS, and 10 ppt for PFNA, PFHxS, and HFPO-DA (GenX). It has also led to increased investment in PFAS treatment technologies and voluntary phase-outs. A Hazard Index approach is also adopted to address combined exposure from PFAS mixtures.118,119
In December 2022, the EPA issued guidance under the National Pollutant Discharge Elimination System (NPDES) framework permitting system for state-level agencies to limit PFAS discharges into water bodies from industrial sources. In August 2022, the EPA proposed designating PFOA and PFOS as hazardous substances under the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA), signaling significant consequences for environmental cleanup initiatives and legal responsibilities regarding liability.
In 2024, the U.S. EPA finalized PFAS reporting and recordkeeping requirements under Toxic Substances Control Act (TSCA) Section 8(a)(7). Entities that have manufactured or imported PFAS or PFAS-containing products since January 1, 2011, must report detailed data on usage, volumes, disposal, exposures, and hazards. However, due to resource constraints, the EPA delayed implementation, with the reporting portal now scheduled to open in July 2025 and close in January 2026.115
Although some of these health effects are well-documented, many studies have yet to establish the connection between PFAS accumulation and damage to various organs, underscoring the need for ongoing research to gain a more comprehensive understanding of potential risks. Through this extensive analysis, the review intends to contribute to a more profound knowledge of the health impacts of short-chain PFAS and provide insights for global regulatory strategies to address their risks effectively. Notably, even within EU member countries and across U.S. states, various agencies have independently enacted distinct policies before adopting more comprehensive union or federal directives.
ADONA | 4,8-Dioxa-3H-perfluorononanoate |
AFFF | Aqueous filmforming foam |
Cl-PFAES | Chlorinated polyfluorinated ether sulfonates |
F53B | Chlorinated polyfluoroalkyl ether sulfonic acid |
CERCLA | Comprehensive environmental response, compensation, and liability act |
EPA | Environmental protection agency |
EU REACH | European union registration, evaluation, authorization, and restriction of chemicals |
FEP | Fluorinated ethylene propylene |
HFPO | Hexafluoropropylene oxide |
HFPO-DA | Hexafluoropropylene oxide dimer acid |
MCLs | Maximum contaminant levels |
NPDES | National pollutant discharge elimination system |
PFAS | Per- and polyfluoroalkyl substances |
PFA | Perfluoroalkoxy alkanes |
PFCs | Perfluoroalkylated compounds |
PFBS | Perfluorobutane sulfonic acid |
PFBS | Perfluorobutanesulfonate |
PFDA | Perfluorododecanoic acid |
PFHpA | Perfluoroheptanoic acid |
PFHxS | Perfluorohexane sulfonate |
PFDA | Perfluoro-n-decanoic acid |
PFNA | Perfluorononanoic acid |
PFOS | Perfluorooctane sulfonic acid |
POSF | Perfluorooctanesulfonyl fluoride |
PFOA | Perfluorooctanoic acid |
PFPE's | Perfluoropolyether |
PFTrDA | Perfluorotridecanoic acid |
PPARα | Peroxisome proliferator-activated receptor alpha |
POPs | Persistent organic pollutants |
PTFE | Polytetrafluoroethylene |
PVDF | Polyvinylidene fluoride |
ROS | Reactive oxygen species |
SHBG | Sex hormone-binding globulin |
SVHC | The substance of very high concern |
TDS | Testicular dysgenesis syndrome |
Supplementary information is available. See DOI: https://doi.org/10.1039/d4va00405a.
This journal is © The Royal Society of Chemistry 2025 |