What is PFAS

PFAS is also known as Perfluoroalkyl and Polyfluoroalkyl substances or “forever chemicals“, are a group of man-made chemicals that includes perfluorooctanoic acid (PFOA), perfluorooctane sulfonic acid (PFOS), perfluorononanoic acid (PFNA), perfluoroalkyl acids (PFAAs), perfluoroalkane sulfonic acids (PFSAs), perfluorohexane sulfonic acid (PFHxS), perfluoroalkyl carboxylic acids (PFCAs), polytetrafluoroethylene (PTFE), also known as Teflon and many others (over 7800 PFAS chemical structures identified to date) ¹, ², ³, ⁴, ⁵, ⁶, ⁷, ⁸. All PFAS contain at least one perfluoroalkyl moiety (C_nF_2n+1–) ⁹. Chemically, individual PFAS can be very different. However, they all have a carbon-fluorine bond, which is very strong and therefore, they do not degrade easily. Fully fluorinated aliphatic carbon chains are known as perfluoroalkyl substances while those with incomplete replacement of hydrogen atoms by fluorine are referred to as polyfluoroalkyl substances.

PFAS have been manufactured and used in the aerospace, automotive, construction, and electronics industries around the world since the 1940s as chemicals that resist grease, oil, water, and heat ¹⁰, ¹¹. PFAS have been widely used for their water resistant (hydrophobic) and oil repellent (oleophobic) properties in consumer products such as disposable food packaging, cookware, outdoor gear, furniture, and carpet ¹², ¹³, ¹⁴, ¹¹. Certain PFAS are also authorized by the US Food and Drug Agency (FDA) for limited use in cookware, food packaging, and food processing equipment ⁵. PFAS are also one of the main components (1–5% w/w) of aqueous film forming foams (fire-fighting foams) used frequently at airports and military bases in firefighting foam for firefighting and training activities ¹⁵, ¹⁶, ¹⁷, ¹⁸. Aqueous film forming foams (fire-fighting foams) contamination of groundwater is a major source of drinking water contamination and has been identified as a nationally significant challenge in countries such as the United States and Sweden ¹⁹, ²⁰. Releases of PFAS to the environment can occur next to chemical manufacturing locations, at industrial sites where PFAS are used, and at various stages of product use and disposal. The carbon-fluorine bond in these compounds is extremely strong and thus many PFAS are not appreciably degraded under environmental conditions ²¹. This has resulted in their accumulation in the environment since the onset of production in the late 1940s ²². In 2015, the Centers for Disease Control and Prevention (CDC) National Health and Nutrition Examination Survey (NHANES) reported that PFAS were found in the blood of nearly all Americans sampled ²³. US companies no longer manufacture the two best-known PFAS, perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) ²⁴. But these legacy PFAS persist in the environment, even as thousands of others remain in production.

People can be exposed to PFAS in different ways, for example through food. Food can become contaminated through contaminated soil and water used to grow the food, through the concentration of these substances in animals via feed and water, through food packaging containing PFAS, or equipment that contained PFAS during food processing. Exposure to PFAS may lead to adverse health effects ²⁵, ²⁶, ²⁷, ²⁸.

Concerns about the public health impact of PFAS have arisen for the following reasons ²⁹:

• Widespread occurrence. Studies find PFAS in the blood and urine of people, and scientists want to know if they cause health problems.
• Numerous exposures. PFAS are used in hundreds of products globally, with many opportunities for human exposure.
• Growing numbers. PFAS are a group of nearly 15,000 synthetic chemicals, according to a chemicals database maintained by the U.S. Environmental Protection Agency (https://comptox.epa.gov/dashboard/chemical-lists/PFASSTRUCT).
• Persistent. PFAS remain in the environment for an unknown amount of time.
• Bioaccumulation. People may encounter different PFAS chemicals in various ways. Over time, people may take in more of the chemicals than they excrete, a process that leads to bioaccumulation in bodies.

Because there are many types of PFAS chemicals, which often occur in complex mixtures and in various everyday products, researchers face challenges in studying them. More research is needed to fully understand all sources of exposure, and if and how they may cause health problems. The research conducted to date reveals possible links between human exposures to PFAS and adverse health outcomes. These health effects include ³⁰, ³¹, ³², ³³, ²:

• altered metabolism,
• fertility,
• reduced fetal growth and increased risk of being overweight or obese,
• increased risk of some cancers,
• reduced ability of the immune system to fight infections.

Potential of PFAS to cause a wide range of negative health impacts depends of various factors, such as the conditions of exposure (dose/concentration, duration, route of exposure, etc.) and characteristics associated with the exposed target (e.g., age, sex, ethnicity, health status, and genetic predisposition) ³⁴. Endocrine disruptive effects of PFAS have been reported to affect fertility, body weight control, thyroid and mammary gland function ³⁵. Developmental effects have been observed in children such as alterations in the behavior or accelerated puberty but also in the newborns such as decreased birthweight ³⁵. Increased risk of kidney, prostate and testicular cancer has been associated with long-term exposure to PFAS in the general population alongside with disturbances in the cholesterol metabolism or reduced efficiency of the immune system against infections ³⁵.

Studies have shown that some PFAS are toxic for both animals and humans ³⁶, ⁶, ⁷, ³⁷, ³⁸, ³⁹, ⁴⁰, ⁴¹. Certain PFAS such as perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS), perfluorononanoic acid (PFNA) and perfluorohexane sulfonic acid (PFHxS), don’t break down in the environment or in the human body, and can accumulate over time. Furthermore, the most frequently detected PFAS, perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA), are highly mobile once introduced to the aquatic environment ⁴² and are not removed by conventional wastewater treatment ⁴³, ⁴⁴. PFAS therefore pose a severe threat to clean water supply worldwide ⁴⁵.

In light of the past and current evidence of PFAS toxic effects, there is an increasing awareness and a general agreement that PFAS substances need to be regulated at multiple levels to minimize their adverse effects on human health and the environment. Manufacturers have phased out production of certain PFAS and in some cases replaced them with new PFAS or chemical substitutes ⁴⁶. For example in textile treatments, many polymers containing long perfluoroalkyl side chains (more than seven perfluorinated carbons) were replaced by analogs containing short perfluoroalkyl side chains (six or four perfluorinated carbons) or fluorine-free moieties (e.g., siloxanes and hydrocarbon polymers) ⁴⁷. Between 2000 to 2002, the main global manufacturer of PFAS (3M Company) voluntarily discontinued manufacturing of the parent chemical perfluorooctane sulfonyl fluoride (POSF) used to produce perfluorooctane sulfonic acid (PFOS) and its precursors perfluorooctane sulfonamide-based chemistry ⁴⁸, ⁴⁹. United States introduced a variety of programs to curb use of the most abundant environmental PFAS, including the PFOA Stewardship Program enacted in 2006 to end production of the longest chained compounds by 2015 ¹¹. Perfluorooctane sulfonic acid (PFOS) was added to the Stockholm Convention’s list of globally restricted Persistent Organic Pollutants (POPs) in 2009 ¹¹. A proliferation of new PFAS have been reported in the environmental literature as industry has rapidly replaced PFOS and PFOA with shorter chain length PFAS and new chemicals that are difficult to detect using standard methods ¹⁶. Emerging evidence from animal experiments suggests some of these alternative PFAS can be equally hazardous ⁵⁰. Environmental health scientists thus face a considerable challenge in understanding the relative importance of diverse exposure pathways to PFAS in different human populations and their potential effects on human health in a rapidly changing chemical landscape ¹¹.

Despite the many regulations to limit the PFAS spread and the phasing out the main PFAS substances (expressed in PFOS and PFOA), other compounds including PFAAs and related substances are still widely used in various industries including fire-fighting foams, photographic, semiconductor and others ⁵¹. Therefore, the detection of PFAS in human bodies have not decreased ⁵¹, ⁵², ⁵³, ⁵⁴.

Figure 1. PFAS chemical structure

Footnote: A general structural formula for perfluoroalkyl substance (PFAS), containing a hydrophobic perfluorinated alkyl tail, and a hydrophilic functional (R) group outlined in a red box. The general structure of non-polymeric, perfluorinated PFAS substances. Non-polymeric PFAS are compounds of variable composition and physicochemical properties that however share two common features. These are represented by the hydrophobic tail, composed by a variable number of carbon atoms at different degree of fluorination, and the hydrophilic head, which contains polar groups. The specific combination of these chemical determinants, namely the carbon chain length, the type of functional groups and the number of fluoride atoms, generates an enormous number of different PFAS molecules with ample downstream applicability. Some of the most common polar groups, are shown.

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Figure 2. PFAS classification

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Figure 3. PFAS substances

Footnote: Examples of non-polymeric and polymeric PFAS substances. (A,B) Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS), two well-known non-polymeric PFAS are shown. Both these compounds possess a relatively long tail containing eight fluorinated carbon atoms, but differ in the chemical composition of the polar head group, which is a carboxylic acid for PFOA and a sulfonic acid for PFOS. Under specific pH environmental conditions, these functional groups can dissociate into the respective anion forms (i.e., carboxylate and sulfonate). (C) Exemplary structure of Polytetrafluoroethylene (PTFE), also known as Teflon, a polymeric PFAS belonging to the subgroup of fluoropolymers. This type of compound is constituted by a moiety of (CF2-CF2)_n atoms which is repeated up to thousands of times; (D) Exemplary structure of a lubrificant also known as Krytox, which is a polymeric PFAS belonging to the subgroup of perfluoropolyethers. In this case the (CF[CF3]−CF2−O)_n moiety is repeated between 10–60 times.

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Figure 4. PFAS emission sources

Footnotes: Schematic illustration of the PFAS environmental distribution and exposure routes for humans and biota. The environmental distribution of PFAS substances involves multiple dispersion routes and exposure pathways that link the sources to the final receptors, represented by humans and wildlife. Industrial manufacture processes, industrial uses and recycling activities, represent primary sources of PFAS emissions. Other indirect sources are the landfilling or the application of contaminated sludge to agriculture land. Volatilization, deposition, leaching and run off processes regulate the redistribution of PFAS between air, soil, water and sediment compartments. Collectively, these pathways contribute to the short-term and long-term exposure of aquatic ecosystems, terrestrial ecosystems and humans to PFAS substances, that can also enter into the food chain through bioaccumulation and indirect human exposure via the ingestion of contaminated food sources.

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Table 1. List of main PFAS categories and subgroups

Non-Polymeric PFAS
Perfluorinated PFAS		Polyfluorinated PFAS
Subgroup	Examples	Subgroup	Examples
Perfluoroalkyl acids (PFAAs) Perfluoroalkane sulfonic acids & sulfonates (PFSAs) Perfluoroalkane sulfnic acids (PFSIAs) Perfluorocarboxylic acids & carboxylates (PFCAs) Perfluoroalkyl phosphonic acids (PFPAs) Perfluoroalkyl phosphinic acids (PFPIAs)	PFBS, PFHxS, PFOS PFOSI PFBA, PFHxA, PFOA C8-PFPA C8/C8-PFPiA	Fluorotelomer compounds (FT)	6:2 FTO, 8:2 FTI
		Perfluoroalkane sulfonamido compounds (Me/Et/Bu-FASAs) Miscellaneous	MeFOSA, FOSE 4,8-Dioxa-3H-perfluorononanoate
Perfluoroalkyl ether acids (PFEAs)	GenX, Adona, F-53B
Perfluoroalkane sulfonamides (FASA)	FOSA
Perfluoroalkane sulfonyl fluorides (PASFs)	PBSF, POSF
Perfluoroalkyl iodides (PFAIs)	PFHxI
Perfluoroalkanoyl fluorides (PAFs)	POF
Perfluoroalkyl aldehydes (PFALs)	PFNAL
Polymeric PFAS
Subgroup	Examples
Fluoropolymers	PVDF, FEP, PFA, ETFE, PTFE (Teflon)
Side-chain Fluorinated Polymers	Fluorinated urethane/acrylate/methacrylate/oxetane plolymers
Perfluoropolyethers (PFPEs)	PFPE-BP, Fluorolink-PFPE

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Table 2. List of PFAS extended chemical names and acronyms

PFAS Chemical Name	PFAS Acronym	PFAS Chemical Name	PFAS Acronym
Perfluorooctanoic acid	PFOA	N-ethyl perfluorooctane sulfonamide	Et-FOSA
Perfluorooctane sulfonic acid	PFOS	N-methyl perfluorooctane sulfonamide	Me-FOSA
Perfluorooctanesulfonamide	PFOSA/FOSA	N-ethyl-perfluorooctane sulfonamido acetic acid	N-Et-FOSAA
Perfluorooctane sulfinic acid	PFOSI	N-methyl-perfluorooctane sulfonamido acetic acid	N-Me-FOSAA
Perfluorononanoic acid	PFNA	2-(N-Methyl-perfluorooctane sulfonamido) acetic acid	Me-FOSAA/Me-PFOSA-AcOH)
Perfluorononanal	PFNAL	2-(N-Ethyl-perfluorooctane sulfonamido) acetic acid	Et-FOSAA/Et-PFOSA-AcOH
Perfluorononane sulfonic acid	PFNS	perfluorooctane sulfonamido ethanol	FOSE
Perfluoroundecanoic acid	PFUnDA	N-ethyl perfluorooctane sulfonamido ethanol	Et-FOSE
Perfluoroundecanoate	PFUnA	perfluorohexane sulfonamide	FHxSA
Perfluorodecanoic acid	PFDA	bis(perfluorooctyl)phosphinic acid	C8/C8-PFPiA
Perfluorododecanoic acid	PFDoA	6:2 Fluorotelomer olefin	6:2 FTO
Perfluorodecane sulfonic acid	PFDS	6:2,8:2,10:2 fluorotelomer alcohol	6:2,8:2,10:2 FTOH
Perfluorobutanoic acid	PFBA	6:2,8:2 fluorotelomer sulfonic acid	6:2,8:2 FTSA
Perfluorobutane sulfonic acid	PFBS	6:2 fluorotelomer thioether amido sulfonate	6:2 FtTAoS
Perfluorophosphonic acid	PFPA	8:2 fluorotelomer iodide	8:2 FTI
Perfluoropentanoic acid	PFPeA	8:2 fluorotelomer unsaturated carboxylic acid	8:2 FTUCA
Perfluoropolyether-benzophenone	PFPE-BP	9-chlorohexadecafluoro-3-oxanone-1-sulfonic acid	6:2 Cl-PFESA (F-53B)
Perfluorohexanoic acid	PFHxA	1-chloroeicosafluoro-3-oxaundecane-1-sulfonic acid	8:2 Cl-PFESA
Perfluorohexyl iodide	PFHxI	Polyvinylidene fluoride	PVDF
Perfluorohexane sulfonic acid	PFHxS	Fluorinated ethylene propylene	FEP
Perfluoroheptanoic acid	PFHpA	Perfluoroalkoxy polymer	PFA
Perfluoroheptane sulfonic acid	PFHpS	Ethylene tetrafluoroethylene	ETFE
Perfluorotridecanoic acid	PFTrDA	Polytetrafluoroethylene	PTFE
Perfluorooctanoyl fluoride	POF	Hexafluoropropylene oxide dimer acid	HFPO-DA/GenX
Perfluorobutane sulfonyl fluoride	PBSF	3H-perfluoro-3-[(3-methoxy-propoxy)] propanoic acid	ADONA
Perfluorooctane sulfonyl fluoride	POSF	Ammonium pentadecafluorooctanoate	APFO
Ammonium perfluoro(2-methyl-3-oxahexanoate)	PMOH

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PFAS Uses

The exceptional strength of the Carbon-Fluorine bond confers very high thermal and chemical stability to PFAS molecules while the presence of hydrophobic and hydrophilic properties along with the variability in the carbon chain length and chemical composition generates an enormous range of different molecules with useful physicochemical properties. For this reason, since 1940’s, PFAS substances have been frequently employed in a vast number of industrial or consumer’s products covering more than 100 sectors of use. Some of the most common PFAS applications include Teflon, pesticide formulation, firefighting foams, surfactants, lubricants, paints, waxes, cosmetics, aerospace, aviation, automotive, textiles coating, water and stain-repellent fabrics, oil production, medical products, food processing, building and construction, energy, paper and packaging, cables and wiring, electronic and semiconductors (see Table 3) ⁵⁵, ⁵⁶, ⁵⁷.

Many PFASs are used as surfactants ⁹. Traditional surfactants comprise a water-soluble hydrophilic portion and a water-insoluble hydrophobic portion. Surfactants lower the surface tension of a liquid, or the interfacial tension between 2 liquids, or between a liquid and a solid. In fluorinated surfactants, the hydrophobic portion contains F bound to C, often as a perfluoroalkyl moiety. The extent of fluorination and location of the F atoms affect the surfactant properties. PFAS surfactants, often referred to as “fluorinated surfactants,” “fluorosurfactants,” “fluorinated tensides,” or “fluorotensides,” are superior in their aqueous surface tension reduction at very low concentrations and are useful as wetting and leveling agents, emulsifiers, foaming agents, or dispersants ¹³, ⁵⁸, ¹⁴. The term “tenside” is encountered most frequently in publications of German origin, and the synonym “surfactant” is preferred in English.

Table 3. PFAS uses in industrial and consumers products

SECTOR OF USE	TYPE OF USE
Non-Polymeric PFAS
Fire prevention	Fire-fighting foams such as foams based on aqueous films (Acqueous Film-Forming Foams, AFFs)
Biocides	Active products in plants grow regulators (PGRs) Active or inert (emulsifiers, solvents, carriers, aerosol propellants) ingredients in pesticides
Electronic	Flame retardants
Aviation and Aerospace	Additives for hydraulic fluids
Metal plating	Humectants and anti-fog agents
Household Products	Surfactants in floor cleaning; treatment for textiles, leather, carpets; car waxes
Building and Construction	Additives in coatings and paints
Medical Products	Stain-resistant and water-repellent articles, X-ray film
Personal care products	Cosmetics, makeup, nail polish, shampoo
Metal plating	Wetting agent, anti-mist agents
Oil and mining production	Surfactants used in oil-well production and mining flotation
PFAS synthesis	Use as monomers for the synthesis of fluoropolymers with fluorinated side chain
Automotive	Treatment for external surfaces and internal leather coatings, textiles or carpets
Textiles and leather	Treatment aimed to create a coating with oil-water-stain-repellent properties
Semiconductors	Use in the production of semiconductor chips
Polymeric PFAS
Fire prevention	Raw materials for firefighting equipment, protective clothes and fuel repellents
Electronic	Insulators and materials for welding
Aviation and Aerospace	Insulators, sleeves
Household Products	Non-stick coatings
Building and Construction	Coating of architectural materials, additives in paints, dyes, stains and sealants
Medical Products	Use in surgical patches, biocompatible human implants and medical prosthesis
Personal care products	Use in dental floss and lotions
Oil and mining production	Use in lining of gas pipes
Automotive	Mechanical components, seals and lubricants
Textiles, leather and clothing	Use in the manufacture of clothing and housewares as well as in coatings having oil-water-repellent properties
Semiconductors	Use as fluids in mechanical vacuum pumps
Energy	Film for solar panels
Paper and packaging	Use in water-oil-repellent materials, paperboard, and bags for food packaging
Cables and wiring	Coatings resistant to weathering, flame and soil
Food processing	Production of materials used for cooking (non-stick pans) and food storage (containers)

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How are humans exposed to PFAS?

The widespread use of PFAS and their persistence in the environment means that PFAS from past and current uses have resulted in increasing levels of contamination of the air, water, and soil. Over time, PFAS may leak into the soil, water, and air. Human exposure to PFAS is widespread and variable by geography and occupation. Because PFAS break down slowly, if at all, people and animals are repeatedly exposed to them, and blood levels of some PFAS can build up over time. People are most likely exposed to PFAS chemicals by consuming PFAS-contaminated water or food, using products made with PFAS, skin contact or breathing air and dust particulate containing PFAS ⁵⁹. Despite the presence of some gaps in our understanding, it is generally accepted that the dietary intake and the consumption of drinking water represent major pathways for the general population ⁶⁰, ⁶¹ while inhalation and dermal contact is far more relevant in case of occupational exposure ⁶², ⁶³, ⁶⁴. In total, 67 to 84 percent of exposure to perfluorooctanoic acid (PFOA) and 88–99% of exposure to perfluorooctyl sulfonic acid (PFOS) in humans come from food products ⁶⁵. Furthermore, the home environment plays a crucial role in total exposure to PFASs. Home air was found to be responsible for 40% of exposure to PFOA in 25% of women who lived in environments with high dust concentrations ⁶⁶.

PFAS food consumption rates vary by age, geography and culture but typical exposure factors are relatively well known ⁶⁷. PFAS concentrations have been reported in milk, meat, vegetables, fruits, and bread in the sub- to low nanogram/g range, while the majority of food samples analyzed contained PFAS below detection limits ⁶⁸, ⁶⁹. In homogenized whole meals, a similar concentration range was reported, although the maximum concentration observed was 118 ng PFOA per gram of fresh food ⁷⁰. Food can become contaminated through contaminated soil and water used to grow the food, through the concentration of these substances in animals via feed and water, through food packaging containing PFAS, or equipment that contained PFAS during food processing. The major exposure pathways for perfluorooctane sulfonic acid (PFOS) for the general population in Europe and North America are food and water ingestion, dust ingestion, and hand-to-mouth transfer from mill-treated carpets ⁷¹. For perfluorooctanoic acid (PFOA), major exposure pathways are oral exposure resulting from migration from paper packaging and wrapping into food, general food and water ingestion, inhalation from impregnated clothes, and dust ingestion ⁷¹. Exposure pathways for other PFAS such as perfluorohexane sulfonic acid (PFHxS) and perfluorononanoic acid (PFNA) are less extensively studied but expected to be similar to perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS). A study reported that PFAS can pass through the placenta to the fetus ⁷². A total of 123 pairs of samples of mothers’ blood and omphalic cord blood were obtained at the Oslo University Hospital. Maternal blood was collected at the 37th week of gestation for the detection of seven PFAS compounds. The level of PFAS measured in cord blood was 30–79% of that of the pregnant mother ⁷².

Since the 1960s, the US Food and Drug Agency (FDA) has authorized specific PFAS for use in specific food contact applications ⁷³. Some PFAS are used in cookware, food packaging, and in food processing for their non-stick and grease, oil, and water-resistant properties. To ensure food contact substances are safe for their intended use, the FDA conducts a rigorous scientific review before they are authorized for the market ⁷³.

PFAS that are authorized for use in contact with food generally fall into four application categories ⁷³:

1. Non-stick cookware: PFAS may be used as a coating to make cookware non-stick.
2. Gaskets, O-Rings, and other parts used in food processing equipment: PFAS may be used as a resin in forming certain parts used in food processing equipment that require chemical and physical durability.
3. Processing aids: PFAS may be used as processing aids for manufacturing other food contact polymers to reduce build-up on manufacturing equipment.
4. Paper or paperboard food packaging: PFAS may be used as grease-proofing agents in fast-food wrappers, microwave popcorn bags, take-out paperboard containers, and pet food bags to prevent oil and grease from foods from leaking through the packaging.

The FDA reviews new scientific information on the authorized uses of food contact substances to ensure that these uses continue to be safe ⁷³. When the FDA identifies potential safety concerns, the agency ensures that these concerns are addressed or that these substances are no longer used in food contact applications. The FDA can work with industry to reach voluntary market phase-out agreements for such food contact substances. The FDA can also revoke food contact authorizations when the agency determines that there is no longer a reasonable certainty of no harm from the authorized use of a food contact substance ⁷³.

One report by the Centers for Disease Control and Prevention, using data from the National Health and Nutrition Examination Survey (NHANES), found PFAS in the blood of 97% of Americans ²³. Another National Health and Nutrition Examination Survey (NHANES) report suggested blood levels of PFOS and PFOA in people have been reduced since those chemicals were removed from consumer products in the early 2000s. However, new PFAS chemicals have been created and exposure to them is difficult to assess. The US Food and Drug Agency (FDA) is undertaking a study of PFAS in food ⁷⁴.

Dietary exposure to PFAS has primarily been estimated using the exposure factor approach by measuring PFAS concentrations in various foods and multiplying by food consumption rates for a given population or demographic group ⁴⁶. A number of studies have estimated dietary exposure to PFAS using the exposure factor approach.

Several studies have used the epidemiologic approach to associate serum PFAS concentrations with different food sources. For example, one study found associations between serum concentrations of several PFAS and fish/seafood consumption in Norway ⁷⁵. In a cohort of 941 American adults with blood sampled between 1996 and 1999, investigators reported positive associations of several PFAS in plasma with consumption of “meat/fish/shellfish (especially fried fish, and excluding omega-3 fatty acid rich fish), low-fiber and high-fat bread/cereal/rice/pasta, and coffee/tea,” but inverse associations with some other foods such as vegetables and fruit ⁷⁶. Another study reported associations between serum perfluorooctanoic acid (PFOA) and perfluorononanoic acid (PFNA) and fast food consumption and take-out coffee in the USA using data from the US National Health and Nutrition Examination Survey (NHANES), suggesting a role for food contact materials ⁷⁷. A different study also based on the US National Health and Nutrition Examination Survey (NHANES) data reported associations between serum PFAS concentrations and fast food restaurant meals as well as microwave popcorn ⁷⁸. A small (n=61) but remarkable Norwegian study examined food consumption using several approaches but report few significant correlations with PFAS in blood ⁷⁹. While diet is likely an important route of exposure for many people, it is difficult to estimate and thus uncertain ⁴⁶. Statistically representative surveys of dietary exposure to PFAS are therefore needed as well as better data on the sources of PFAS found in food and links to those present in food contact materials ⁴⁶.

Human exposures to perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) have been declining in western countries and Japan over the last decade due to these regulatory interventions because of their adverse effects on human health ⁸⁰, ⁸¹, ⁸², ⁸³, ⁸⁴.

PFAS Health Effects

While the science surrounding potential health effects of PFAS is developing, exposure to some types of PFAS have been associated with serious health effects such as liver toxicity, reproductive disorders, neurotoxicity and immunotoxicity ⁸⁵. PFAS are potentially capable of producing a wide range of adverse health effects depending on various factors, such as the conditions of exposure (dose/concentration, duration, route of exposure, etc.) and factors associated with the individuals exposed (e.g., age, sex, ethnicity, health status, and genetic predisposition) ³⁴. Moreover, the mechanisms underlying toxic health effects attributed to PFAS in humans are not well understood ³. There is strong evidence that PFAS toxic effects that are observed in rodents include liver damage (hepatotoxicity), damage to the immune system (immunotoxicity) and developmental toxicity (e.g., increased fetal and/or neonatal mortality and reduction in fetal weight and/or postnatal growth), involve the activation of peroxisome proliferator-activated receptor-alpha (PPARα); however, humans and nonhuman primates are less responsive to peroxisome proliferator-activated receptor-alpha (PPARα) than rodents ⁸⁶, ⁸⁰. In addition, peroxisome proliferator-activated receptor-alpha (PPARα)-independent mechanisms are also involved, and it is not known if species differences exist for these mechanisms ⁴.

The available epidemiology studies suggest the following associations between PFAS exposure and several adverse health effects ⁶, ⁷, ³⁷, ³⁸, ³⁹, ⁴⁰, ⁴¹:

• Increased cholesterol levels (PFOA, PFOS, PFNA);
• Changes in liver enzymes (PFOA, PFOS, PFHxS);
• Decreased vaccine response in children (PFOA, PFOS, PFHxS);
• Increased risk of high blood pressure or preeclampsia in pregnant women (PFOA, PFOS); and
• Decreases in infant birth weights (<20 g (0.7 ounces) fall in birth weight per 1 ng/ml elevation in PFOA or PFOS in blood).

Selected human studies (published since 2010) investigating toxic effects of PFAS are presented in Table 4. Majority of the human studies explored the linkage between PFAS concentration and lipid status, mainly cholesterol level ⁸⁷, ⁸⁸, while a study was also conducted to assess the connection between PFAS and cholesterol at the gene expression level ⁸⁹. Multiple studies find significant associations between PFAS exposure and adverse immune outcomes in children ¹¹. Abnormally elevated cholesterol or fats in the blood (dyslipidemia) is the strongest metabolic outcome associated with PFAS exposure. Eriksen et al. ⁸⁷ discovered substantial positive relationships between PFOS, PFAS, and total cholesterol in 753 individuals, while sex and prevalence of diabetes were suggested to influence the connection between these two substances and cholesterol. Fletcher et al. ⁸⁹ observed an inverse relationship between serum PFOA levels and the expression level of genes involved in cholesterol transport in whole blood (NR1H2, NPC1 and ABCG1). A positive correlation was found between PFOS and a transcript involved in cholesterol mobilization (NCEH1), while a negative relationship was seen between PFOS and a transcript involved in cholesterol transport (NCEH2). Sex-specific effects were also noticed in Fletcher et al. study ⁸⁹. On the other hand, in a study involving 815 participants ≤18 years of age, Geiger et al. ⁸⁸ found that serum PFOA and PFOS were related with high total cholesterol and LDL-C levels, regardless of age, gender, race-ethnicity, body mass index, yearly family income, physical activity, or serum cotinine levels. PFOA and PFOS were not shown to be substantially linked with aberrant HDL-C and triglyceride levels.

Other studies explored the link between PFAS concentration and different hormones, such as thyroid ⁹⁰, ⁹¹ and sex hormones ⁹¹, ⁹², ⁹³, as well as development ⁹⁴, ⁹². By assessing the connection between the levels of 14 PFAS in healthy men from the general population and different sex hormones and semen sample quality, Joensen et al. ⁹⁵ found that only PFOS levels were negatively associated with testosterone, calculated free testosterone (FT), free androgen index (FAI) and ratios of T/LH, FAI/LH and FT/LH. Other PFAS were found at lower levels than PFOS and did not exhibit the same associations ⁹⁵. Also, after measuring PFAS levels in 1682 males and females 12 to 80 years of age, Lewis et al. ⁹⁰ found no significant relationships between any of the PFAS and testosterone. PFAS were suggested to be associated with increases in FT3, TT3, and FT4 among adult females. The authors concluded that, during the adolescence, PFAS may be related to increases in TSH among males and decreases in TSH among females, suggesting sex-specific effects ⁹⁰. In contrast, Li et al. ⁹¹ discovered no associations between PFAS and thyroid hormones in adults and seniors in 3297 participants from Ronneby, a municipality with highly contaminated drinking water by PFAS (exposed group), with the exception of a positive association between PFAS and fT4 in males over 50. Thyroid hormone levels were observed to be higher in Ronneby preteen children compared to the control group. Weak evidence of a link between increasing PFAS levels and lower fT3 in preteen boys and lower TSH in adolescent men was found ⁹¹.

Lopez-Espinosa et al. ⁹⁴ aimed to investigate whether PFOS and PFOA were linked to the markers of sexual development. Their study included 3076 boys and 2931 girls aged 8–18 years. For boys, there was a link between increased PFOS and a lower chance of reaching puberty. Higher PFOA or PFOS concentrations in girls were related to a lower risk of post menarche ⁹⁴. The same group of researchers examined the link between PFAS levels and estradiol, total testosterone, and IGF-1 in 2292 children. In boys, PFOA concentrations were substantially related to testosterone levels; PFOS concentrations were related to estradiol, testosterone, and IGF-1, while PFNA concentrations were linked to IGF-1. Significant linkage was discovered in girls between PFOS and testosterone and IGF-1, as well as PFNA and IGF-1 ⁹⁶. Furthermore, Wang et al. ⁹² concluded that PFAS may affect estrogen homeostasis and foetal growth during pregnancy, and that estrogens may mediate the relationship between PFAS exposure and foetal growth after examining 424 mother-infant pairs.

Some of the studies also explored the linkage between the PFAS exposure and liver function ⁹⁷, ⁹⁸, ⁹⁹. In 47,092 adult participants, Gallo et al. ¹⁰⁰ found a positive association between PFOA and PFOS concentrations and serum ALT level. On the other hand, the relationship with bilirubin appeared to increase at low levels of PFOA and decrease at higher levels ¹⁰⁰. In 1002 individuals from Sweden, Salihovic et al. ¹⁰¹ have also found a positive association of PFHpA, PFOA, PFNA, and PFOS concentrations and ALT activity, but also positive associations of PFHpA, PFOA, PFNA, PFDA, and PFUnDA with ALP. These authors noted that the changes of investigated PFAS concentrations were positively associated with gamma glutamyl transferase (GGT) levels and inversely associated with the changes in circulating bilirubin ¹⁰¹. On the other hand, in 2883 participants, (1801 non-obese and 1082 obese), Jain and Ducatman ¹⁰² investigated the connection between liver function alterations and various PFAS. They concluded that connections might only be observed in the obese participants: alanine aminotransferase (ALT) was positively associated with PFOA, PFHxS, and PFNA. On the other hand, PFOA and PFNA were associated with GGT ¹⁰². Epidemiological studies revealed a connection between PFAS and decrease in vaccination antibody production in early infants and children, especially having in mind that, if breastfed, they have a relatively high exposure and may be more susceptible as their immune system develops. Abraham et al. ⁹³ found significant associations between the concentration of PFOA, but not PFOS, and adjusted levels of vaccine antibodies against Haemophilus influenza type b, tetanus and diphtheria for which no observed adverse effect concentrations (NOAECs) were 12.2, 16.9 and 16.2 µg/L, respectively. Furthermore, PFOA levels were shown to be inversely related to the interferon gamma (IFN-γ) production of ex-vivo lymphocytes after stimulation with tetanus and diphtheria toxoid ⁹³. Furthermore, Budtz-Jorgensen E and Grandjean P ¹⁰³ found an approximate BMDL of 1 ng/mL serum for both PFOS and PFOA for the serum concentrations of specific IgG antibodies against tetanus and diphtheria at ages 5 and 7 as outcome parameters. These authors proposed the reference concentration of about 0.1 ng/mL as the serum-based target, a level which is below the most reported human serum-PFAS concentrations ¹⁰³. Grandjean et al. ¹⁰⁴ discovered that prenatal exposure to PFAS had an inverse relationship with antibody concentrations five years later, while concentrations measured at 3 and 6 months of age had the highest inverse relationships with antibody concentrations at 5 years of age, especially for tetanus. The same authors have found that diphtheria antibody concentrations dropped at higher PFAS concentrations at 13 and 7 years after booster vaccinations at 5 years of age; the correlations were statistically significant for PFDA at 7 years and PFOA at 13 years, implying a 25% decrease for each doubling of exposure ⁹⁸.

Evidence for cancer is limited to manufacturing locations with extremely high exposures and insufficient data are available to characterize impacts of PFAS exposures on neurodevelopment ¹¹. The International Agency for Research on Cancer concluded that perfluorooctanoic acid (PFOA) is possibly carcinogenic to humans (Group 2B) and the United States Environmental Protection Agency indicated that there was suggestive evidence of cancer causing potential of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) in humans ¹⁰⁵, ¹⁰⁶, ¹⁰⁷, ¹⁰⁸. Increases in testicular and kidney cancer were noted in highly exposed humans ¹⁰⁹, ¹¹⁰, ¹¹¹.

Figure 5. PFAS effects on human health

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Table 4. Selected human studies exploring the toxicity of PFAS

Substance	Population	Measured Parameters	Results	References
PFOS PFOA	middle-aged Danish population; 753 individuals (663 men and 90 women), 50–65 years of age, nested within a Danish cohort of 57,053 participants	serum levels of total cholesterol	Statistically significant positive associations between PFOS, PFAS and total cholesterol level Sex and prevalent diabetes modified the association between PFOA and PFOS and cholesterol	87
PFOS PFOA	815 participants ≤18 years of age from the National Health and Nutrition Examination Survey 1999–2008	dyslipidemia: total cholesterol >170 mg/dL, low-density lipoprotein cholesterol (LDL-C) >110 mg/dL, high-density lipoprotein cholesterol (HDL-C) <40 mg/dL or triglycerides >150 mg/dL.	Serum PFOA and PFOS-positively associated with high total cholesterol and LDL-C, independent of age, sex, race-ethnicity, body mass index, annual household income, physical activity and serum cotinine levels PFOA and PFOS-not significantly associated with abnormal HDL-C and triglyceride levels.	88
PFOS PFOA	290 individuals (144 men + 146 women) exposed to background levels of PFOS and elevated concentrations of PFOA through drinking water, aged between 20 and 60 years	expression of genes involved in cholesterol metabolism	Inverse associations between serum PFOA levels and the whole blood expression level of genes involved in cholesterol transport (NR1H2, NPC1 and ABCG1) A positive association between PFOS and a transcript involved in cholesterol mobilisation (NCEH1), and a negative relationship with a transcript involved in cholesterol transport (NR1H3) Reductions in the levels of mRNAs involved in cholesterol transport were seen with PFOA in men (NPC1, ABCG1, and PPARA) and in women (NR1H2 expression) Increase in the levels of a cholesterol mobilisation transcript (NCEH1) in women. PFOS was positively associated with expression of genes involved in both cholesterol mobilisation and transport in women (NCEH1 and PPARA)	89
PFOA PFOS PFHxS PFNA PFDA	2883 participants, (1801 non-obese and 1082 obese), aged more than or equal to 20 years old	liver function parameters: AST, ALT, GGT, ALP, and total bilirubin (TB)	Among obese participants only, alanine aminotransferase (ALT)-positively associated with PFOA, PFHxS, and PFNA PFOA and PFNA were associated with gamma GGT in obese participants	102
14 PFCs	Healthy men from the general population, median age of 19 years	total testosterone (T), estradiol (E), sex hormone-binding globulin (SHBG), luteinizing hormone (LH), follicle-stimulating hormone (FSH) and inhibin-B and Semen samples analysis	PFOS levels-negatively associated with testosterone, calculated free testosterone (FT), free androgen index (FAI) and ratios of T/LH, FAI/LH and FT/LH Other PFCs were found at lower levels than PFOS and did not exhibit the same associations. PFC levels were not significantly associated with semen quality	95
PFOA PFOS PFHxS PFNA	1682 males and females 12 to 80 years of age	testosterone (T), thyroid stimulating hormone (TSH), and free and total triiodothyronine (FT3, TT3) and thyroxine (FT4, TT4)	Exposure to PFAS may be associated with increases in FT3, TT3, and FT4 among adult females During adolescence, PFAS may be related to increases in TSH among males and decreases in TSH among females No significant relationships were observed between PFAS and T in any of the models	90
PFOS PFOA	3076 boys and 2931 girls aged 8–18 years	subjects were classified as having reached puberty based on either hormone levels (total >50 ng/dL and free >5 pg/mL testosterone in boys and estradiol >20 pg/mL in girls) or onset of menarche	For boys, there was a relationship of reduced odds of reached puberty (raised testosterone) with increasing PFOS (delay of 190 days between the highest and lowest quartile) For girls, higher concentrations of PFOA or PFOS were associated with reduced odds of postmenarche (130 and 138 days of delay, respectively)	94
PFOS PFOA PFNA	2292 children (6–9 years of age)	estradiol, total testosterone, and IGF-1	In boys, PFOA concentrations were significantly associated with testosterone levels; PFOS with estradiol, testosterone, and IGF-1; and PFNA with IGF-1 In girls, significant associations were found between PFOS and testosterone and IGF-1; and PFNA and IGF-1	96
PFOS PFOA	424 mother-infant pairs	estrone (E1), b-estradiol (E2), and estriol (E3), infants: head circumference, body weight, body length	PFOS was positively related to E1 and E3, but negatively related to E2 Serum PFOA was positively related to serum E1 and negatively related to head circumference at birth Serum E2 was negatively related to head circumference, body weight, and body length at birth and serum E3 was positively related to body weight Serum E3 mediated the relationship between serum PFOS and body weight PFAS could affect estrogen homeostasis and fetal growth during pregnancy and estrogens might mediate the association between exposure to PFAS and fetal growth	92
PFOS PFOA	47,092 adults	alanine transaminase (ALT), γ-glutamyltransferase (GGT), direct bilirubin	Positive association between PFOA and PFOS concentrations and serum ALT level, a marker of hepatocellular damage. The relationship with bilirubin appears to rise at low levels of PFOA and to fall again at higher levels.	100
PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFHxS PFOSA	1002 individuals from Sweden (50% women) at ages 70, 75 and 80	bilirubin and hepatic enzymes alanine aminotransferase (ALT), alkaline phosphatase (ALP), and γ-glutamyltransferase (GGT)	Positive associations of PFHpA, PFOA, PFNA, PFDA, and PFUnDA with ALP Concentrations of PFHpA, PFOA, PFNA, and PFOS were positively associated with the activity of ALT The changes in PFAS concentrations were positively associated with GGT and inversely associated with the changes in circulating bilirubin	101
PFOS PFOA PFHxS	3297 participants from Ronneby, a municipality with drinking water highly contaminated by PFAS (exposed group)	thyroid hormone levels, with adjustments for age, sex and BMI	No associations between PFAS and thyroid hormones in adults and seniors except for a positive association between PFAS and fT4 in males over 50 Higher thyroid hormone levels in the preteen children from Ronneby compared to the reference group Weak evidence of associations between increased PFAS levels and decreased fT3 in preteen boys, and decreased TSH in teenage males	112
PFOA PFOS	101 healthy 1-year-old children	Antibodies against haemophilus infuenza type b, tetanus and diphtheria, interferon gamma, cholesterol	Significant associations between PFOA, but not PFOS concentrations, and adjusted levels of vaccine antibodies against haemophilus influenza type b, tetanus and diphtheria PFOA levels inversely related to the interferon gamma (IFN) production of ex-vivo lymphocytes after stimulation with tetanus and diphtheria toxoid No infuence of PFOA and PFOS on infections and cholesterol level during the frst year of life	93
PFOA PFOS	1146 children	serum concentrations of specific IgG antibodies against tetanus and diphtheria at ages 5 and 7	Approximate BMDL of 1 ng/mL serum for both PFOS and PFOA for the serum concentrations of specific IgG antibodies against tetanus and diphtheria at ages 5 and 7 Proposed reference concentration of about 0.1 ng/mL as the serum-based target	103
PFHxS, PFOS, PFOA, PFDA, PFNA	275 males and 349 females participated in clinical examinations and provided blood samples at ages 18 months and 5 years	serum concentrations of antibodies against tetanus and diphtheria vaccines determined at age 5	Pre-natal exposure showed inverse associations with the antibody concentrations five years later, with decreases by up to about 20% for each two-fold higher exposure Associations for serum concentrations at 18 months and 5 years were weaker Concentrations estimated for ages 3 and 6 months showed the strongest inverse associations with antibody concentrations at age 5 years, particularly for tetanus Joint analyses showed statistically significant decreases in tetanus antibody concentrations by 19–29% at age 5 for each doubling of the PFAS exposure in early infancy	104
PFHxS, PFOA, PFOS, PFNA, PFDA.	516 subjects	PFAS serum concentrations and concentration of antibodies against diphtheria and tetanus	Diphtheria antibody concentrations decreased at elevated PFAS concentrations at 13 y and 7 y; the associations were statistically significant for perfluorodecanoate (PFDA) at 7 y and for perfluorooctanoate (PFOA) at 13 y, both suggesting a decrease by ∼25% for each doubling of exposure Structural equation models showed that a doubling in PFAS exposure at 7 y was associated with losses in diphtheria antibody concentrations at 13 y of 10–30% for the five PFAS	98

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PFAS Cholesterol and Metabolic effects

Cross-sectional and longitudinal studies published between 1996–2020 based on human populations in Australia, Canada, China, several European countries, Japan, South Korea, Taiwan, UK and the U.S. found consistent evidence of modest positive associations between elevated serum PFAS concentrations and detrimental lipid profiles, such as elevated total cholesterol and low-density lipoprotein cholesterol (LDL-C), or reduced high-density lipoprotein cholesterol (HDL-C) in adults and children, although the magnitude of the cholesterol effect is inconsistent across different exposure levels ¹¹, ¹¹³, ¹¹⁴, ¹¹⁵, ¹¹⁶, ¹¹⁷, ¹¹⁸, ¹¹⁹, ¹²⁰, ¹²¹, ¹²², ¹²³, ¹²⁴, ¹²⁵, ¹²⁶, ¹²⁷, ¹²⁸, ¹²⁹, ¹³⁰, ¹³¹. Perfluorooctanesulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) exhibit the most consistent finding across studies. The effect size varies across studies, which can be a result of different exposure levels. Increases in serum PFOA and PFOS from the lowest to the highest quintiles among children in C8 health project was associated with 4.6 and 8.5 mg/dL total cholesterol (reference level for children is <170 mg/dL) ¹³¹. Among the National Health and Nutrition Examination Survey (NHANES) 2003–2004 participants, increases in serum PFOA and PFOS from the lowest to the highest quartiles were associated with 9.8 and 13.4 mg/dL total cholesterol (reference level for adults is <200 mg/dL) ¹¹⁴. Maisonet et al. ¹³² reported a non-linear relationship between prenatal PFOA concentrations and total cholesterol at ages 7 and 15 of the child.

Studies of large populations, featuring wide exposure ranges, demonstrate that serum lipids rapidly increase beginning at background PFAS (1–10 ng/mL) serum concentration and then are followed by attenuating (“plateaued”) cholesterol measurements as (log-transformed) exposures to long-chain PFAS increase ¹²⁶, ¹¹³, ¹³¹. These findings suggest partially saturable mechanisms; thus, the cholesterol dose response at pharmacologic or acutely toxic doses should be viewed with caution; associations can be missed or may be misleading when an environmental range of exposure is absent. At background PFAS exposure levels, residual associations may be more detectable in obese participants ¹³³, ¹³⁴, a finding congruent with experimental PFAS outcomes in rodents fed “Western” or high-fat diets laced with PFAS ¹³⁵, ¹³⁶, ¹³⁷. Human gene expression pathways provide support for an interaction of obesity and PFAS exposures and suggest possible sex differences ¹³⁸. A pharmacokinetic model predicts that approximately half of the PFOS-exposed population would experience a >20% rise in serum cholesterol ¹³⁹. Risk-assessment implications for low-PFAS dose increases in cholesterol have been noted ¹⁴⁰, ¹²⁶ and a review of population and toxicity data concluded that dyslipidemia is the strongest metabolic outcome of PFAS exposure ¹¹.

Human PFAS lipid findings may be related to experimental findings of induced adipogenesis, impaired bile acid metabolism/synthesis, strongly decreased CYP7A1 enzyme activity, altered fatty acid transport, and intracellular lipid accumulation with steatosis, including in PPAR-α-null or PPAR-α-humanized animals ¹⁴¹, ¹⁴², ¹⁴³, ¹⁴⁴, ¹⁴⁵, ¹⁴⁶, ¹⁴⁷, ¹⁴⁸, ¹⁴⁹, ¹⁵⁰, ¹⁵¹. Independent of PFAS exposure, similar alterations in metabolic pathways have been related to disrupted fatty acid beta-oxidation and increased free cholesterol in toxicology studies ¹⁵².

Emerging longitudinal and diabetes clinical trial data indicate that PFAS may increase human insulin resistance, associated with dysregulated lipogenesis activity ¹⁵³, ¹²⁷. Longitudinal studies of clinically diagnosed diabetes patients have sometimes associated PFAS exposures with diabetes ¹⁵⁴ or with small changes in glycemic markers ¹⁵⁵; however, diabetes associations to date are not consistent ¹⁵⁶, ¹⁵⁷, ¹⁵⁵. Future studies should consider whether PFAS may instigate autoimmune diabetic outcomes in humans, as shown in experimental studies ¹⁵⁸. Experimental data reveal that PFAS activate G protein-coupled receptor 40, a free fatty acid-regulated membrane receptor on islet ß cells, stimulating insulin secretion ¹⁵⁹, ¹⁶⁰.

There is some but much less consistent evidence of a modest positive correlation with metabolic diseases such as diabetes, overweight, obesity and heart diseases ¹¹. The majority of studies are cross-sectional, which have limited causal interpretation ¹⁶¹. A few studies provided stronger evidence than observational studies, such as Diabetes Prevention Program Trial where at baseline, a doubling in plasma perfluorooctanesulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) concentrations was associated with small differences in markers of insulin secretion and β – cell function. However, there was limited evidence suggesting that PFAS concentrations are associated with diabetes incidence or changes in glycemic indicators during the follow-up period ¹⁵⁵. And in the diet-induced weight-loss trial, higher baseline plasma PFAS concentrations were associated with a greater weight regain, especially in women ³⁰.

An odds ratio (OR) represents the odds that an outcome will occur given a particular exposure, compared to the odds of the outcome occurring in the absence of that exposure ¹⁶²:

• Odds Ratio (OR) = 1 Exposure does not affect odds of outcome
• Odds Ratio (OR) greater than 1 means Exposure is associated with higher odds of outcome
• Odds Ratio (OR) less than 1 means Exposure is associated with lower odds of outcome

Eighteen studies have examined the associations between PFAS exposures and glucose metabolism, insulin resistance and diabetes. Overall the results across different studies are inconclusive ¹¹. Lin et al. ¹⁶³ was the first to report a positive association between serum PFAS concentrations and glucose homeostasis among adults and adolescents in the National Health and Nutrition Examination Survey (NHANES). They reported a considerable effective size – doubling serum PFNA concentrations was associated with hyperglycemia odds ratio (OR) of 3.16 ¹⁶³. Later studies tend to report smaller effect sizes. Exposure during pregnancy may affect the mother and child during gestation and later in life. In a small pregnancy cohort in the U.S., each standard deviation of increase in PFOA was associated with a 1.87-fold increase of gestational diabetes risk ¹⁶⁴. In a larger Spanish cohort ¹⁶⁵, a null result was reported for PFOA, but PFOS, PFHxS and gestational diabetes had positive associations: Odds Ratio (OR) per log10-unit increase=1.99 and Odds Ratio (OR)=1.65, respectively.

Results for hypertension and other vascular diseases including stroke are also inconsistent ¹¹. Two of the earliest studies examined the relationship between PFAS exposure and hypertension among NHANES and found different results for children and adults. Adjusted Odds Ratio (OR)=2.62 for hypertension comparing 80th vs. 20th percentiles serum PFOA among NHANES adults in the U.S. ¹⁶⁶, while among children a null finding was reported ¹⁶⁷. In some later cohort studies, null results and even protective effects associated with PFAS exposure and hypertension were reported ¹⁶⁸, ¹⁶⁹. A cross-sectional study on carotid artery intima-media thickness in adolescents reported increased risks with increase in plasma perfluorooctanesulfonic acid (PFOS) ¹⁷⁰. However, a more recent study on artery stiffness found protective effects of PFOA and PFNA among children and adolescents enrolled in the World Trade Center Health Registry ¹²⁴.

Other metabolic endpoints include thyroid disease (which could also be considered an endpoint for endocrine disruption), cardiovascular diseases, uric acid metabolism, and body weight. Except for uric acid metabolism, most results are inconclusive ¹¹. An increase in hyperuricemia risks and PFOA exposure was observed in all four studies (two from NHANES and two from C8 Health Project) ¹¹.

In summary, the strongest evidence for a relationship between PFAS exposure and metabolic outcome is in the area of dyslipidemia ¹¹. Animal studies have found decreases in serum cholesterol levels associated with increased PFAS exposures, which contradicts epidemiological findings ¹¹. The difference may lie in different levels of expression for nuclear receptors involved in the toxicological pathway, such as peroxisome proliferator-activated receptor (PPAR)-alpha. It may also be related to differences in exposure levels. Dietary factors can influence metabolic outcomes ¹⁷¹, introducing bias into observed relationships if not controlled for properly. Explanations for null findings include healthy worker effects and non-linear relationships, such as a decreasing slopes as exposure increases (log-linear relationships) ¹⁷².

PFAS Liver effects

The liver is a primary target organ for long-chain PFAS storage, and accompanying experimental evidence of toxicity includes hepatocyte fat infiltration, specific P450 (CYP) pathway induction, apoptosis, hepatocellular adenomas and carcinomas, and disrupted fatty acid trafficking that can be peroxisome proliferator-activated receptor alpha (PPARα)-dependent or -independent and present across species ¹⁷³, ¹⁷⁴, ¹⁷⁵, ¹⁷⁶, ¹⁷⁷, ¹⁷⁸, ¹⁴³, ¹⁷⁹.

Population studies demonstrate significant associations of long-chain PFAS (>6 fluorinated carbons) exposure to higher liver enzymes, such as alanine aminotransferase in adults and adolescents ¹⁸⁰, ¹⁸¹, ¹⁸², ¹⁸³, ¹⁸⁴, ¹⁸⁵, including in longitudinal studies ¹⁸⁶, ¹⁸⁷. Following low-dose exposures, these associations may be more evident in obese participants ¹⁷², ¹⁸¹, ¹⁸⁸.

Based on experimental data, nonalcoholic fatty liver disease (NAFLD) has been investigated as a clinical outcome of PFAS exposure mediating consistent population PFAS-altered liver enzyme findings ¹⁸⁹, ¹⁹⁰, ¹⁹¹, ¹⁴⁵. Studies with nonalcoholic fatty liver disease (NAFLD) cytokeratin C18 biomarkers have provided supportive evidence for PFAS inducing steatosis ¹⁹². Metabolomic studies have been directed at potentially explanatory human glycerophosphocholine and fatty acid profiles ¹⁹³, ¹⁴⁴, ¹⁹⁴. Processes which favor steatosis promote advanced liver disease including liver cancer in humans ¹⁹⁵. Associations of PFAS with advanced human liver disease and liver cancer are technically hard to study for reasons including (and not limited to) lethality, selection of comparison populations, and alterations of excretion mechanics associated with disease states. In a clinic-based study, mostly obese (85%) children aged 7 to 19 years with biopsy-proven NAFLD had more advanced disease associated with PFOS and PFHxS exposure as well as associations with lipid and amino acid pathways linked to NAFLD pathogenesis ¹⁹⁶. However, an adult study reported that serum PFHxS was inversely associated with hepatic lobular inflammation in morbidly obese bariatric surgery patients ¹⁹⁷. A study of heavily exposed workers (n = 462, geometric mean serum PFOA of 4048 ng/mL) detected significantly increased incident mortality for cirrhosis and liver cancer compared to a regional population ¹⁹⁸, whereas no PFAS association to cancer or advanced liver disease was reported in a 3M worker cohort or in the C8 Health study population ¹⁹⁹, ²⁰⁰, ²⁰¹.

Emerging animal toxicology and histology and human population data provide mechanistic clues that PFAS disrupt hepatic metabolism, leading to increased bile acid reuptake and lipid accumulation in liver ²⁰², ¹⁴¹. A review of NAFLD and toxicant exposure concluded that PFAS are associated with early steatosis (“fatty liver”), the preclinical stage of NAFLD ²⁰³.

PFAS Neurodevelopmental effects

In vitro studies suggest perfluorooctanesulfonic acid (PFOS) can trigger the “opening” of tight junction in brain endothelial cells and increase the permeability of the blood brain barrier ²⁰⁴. There has therefore been some interest in investigating the neurotoxic effects associated with PFAS exposures. In laboratory animals, it has been reported that PFOS, PFOA and PFHxS exposures during the peak time of rapid brain growth in mice resulted in an inability to habituate in unfamiliar environment ²⁰⁵. A few studies reported neurotoxicity of PFOS, PFHxS, and PFOA in cell culture systems ²⁰⁶, as well as altered behavioral responses ²⁰⁷ and deficits in learning and memory ability in rodents ²⁰⁸. In contrast, no significant developmental neurotoxic effects were seen from prenatal exposure to PFOS in US EPA guideline-based studies with rats ²⁰⁹.

Liew et al ²¹⁰ reviewed 21 epidemiological studies in 2018 and concluded that evidence is mixed regarding neurodevelopmental effects of PFAS exposures. Health outcomes examined included developmental milestones in infancy, attention-deficit/hyperactivity disorder (ADHD) and behaviors in childhood, and neuropsychological functions such as IQ and other scales or scores ²¹⁰. Neurodevelopmental trajectories are highly complicated and there is great heterogeneity in the instruments and methods to evaluate neurodevelopmental endpoints. Existing evidence suggests that PFAS can impact the nervous system, with particularly harmful effects from developmental exposures or exposures in sensitive populations ²¹¹. However, the limitations and inconsistencies in the current research make the severity of the neurotoxicological ramifications of PFAS exposure largely unknown ²¹¹. Additional research is needed to establish a link between neurodevelopmental outcomes and PFAS exposures.

PFAS Immune effects

Immunotoxicity of PFASs has been demonstrated in multiple animal models, including rodents, birds, reptiles and other mammalian and non-mammalian wildlife. Epidemiological data is relatively sparse but mounting evidence suggests that the immunotoxic effects in laboratory animal models occur at serum concentrations that are comparable to body burden of highly exposed humans and wildlife ²¹².

The health outcomes related to PFAS immunotoxicity include both molecular-level (i.e. antibody concentrations) and organ/system-level (i.e. infection of respiratory system). In general, more consistent results across different studies were reported for molecular-level health endpoints such as vaccine antibody or other immune markers such as immunoglobulin ¹¹.

Five studies examined the association between PFAS exposure and suppression of antibody response to vaccination among children, adolescents or adults. Four out of the five found statistically significant associations between higher PFAS exposure and suppressed immune response ¹¹. Grandjean et al. ²¹³ was the first to link PFAS exposure in children to deficits in immune function. The authors reported a 2-fold increase in perfluorooctane sulfonate (PFOS) in child serum was associated with a −49% decline in tetanus and diphtheria antibody concentrations ²¹³. Decreased immunological response persisted at age 13 years ²¹⁴. This effect size is larger than later studies and can be attributed to different exposure levels, different vaccine strains, and different times elapsed since vaccination (peak antibodies vs residual antibodies). Adverse associations have also been found in PFAS exposure and other childhood vaccinations such as rubella, mumps, and Hemophilus influenza vaccinations in children ²¹⁵, ²¹⁶, ²¹⁷ and adult influenza vaccination such as FluMist ²¹⁸ and anti-H3N2 ²¹⁹. In a single study, modest down-regulation of C-reactive protein response, a marker of human systemic inflammation, was also reported to be associated with perfluorooctanoic acid (PFOA) blood levels ²²⁰.

Disease outcomes linked with immunosuppression such as clinician-recorded diagnoses of childhood infections have also been associated with prenatal exposures to PFOS and perfluorohexane sulfonate (PFHxS) ²²¹. A pregnancy cohort study prospectively detected increased risk of airway and throat infections and diarrhea in children through age 10 yr, correlated with cord-blood PFAS measurements ²²². A recent review concluded that exposure to PFAS in infancy and childhood resulted in an immunosuppressive effect characterized by an increased incidence of atopic dermatitis and lower respiratory tract infections ²²³. Some of the immunological effects were sex-specific, but the authors cautioned that there were inconsistencies across studies ²²³. Overall, available data provide strong evidence that PFAS exposure can suppress the human immune response.

Population studies of immune hyperreactive diseases have resulted in mixed findings. Studies on childhood allergy and asthma outcomes have shown no association with PFAS ²²⁴, ²²², whereas others have found substantial effects, including provocative evidence that subgroups of individuals not adequately immunized may be at an increased risk for disease a priori ²²⁵, ²²⁶. For example, a case-control study of Taiwanese children compared the first and fourth quartiles of serum measurements for 11 PFAS with asthma and other immune markers and reported confidence intervals well above 1.0 for PFOA and others ²²⁵. However, review articles concerning PFAS and childhood allergy and asthma offer nuanced, age- and sex-specific interpretations and advise against firm conclusions ²²³.

Chronic autoimmune outcomes, including thyroid disease and inflammatory bowel disease (IBD), have also been considered. A study in contaminated communities (n = 32 254) detected an association between both prevalence and incidence of ulcerative colitis (UC) and PFOA exposure ²²⁷. A worker study (n = 3713) found a higher prevalence and incidence of ulcerative colitis (UC) with increasing log PFOA serum concentrations ¹⁶⁹. A case-control study of children and young adults from a background exposure community in Atlanta, Georgia, USA, also found higher serum PFOA levels in patients with ulcerative colitis (UC) ²²⁸. In contrast to PFOA-related associations in US populations, a study of a contaminated community in Sweden (n = 63 074) did not show a consistent association of inflammatory bowel disease (IBD) with any PFAS exposure ²²⁹.

Recent, thorough reviews emphasize the following key concepts ²³⁰, ²³¹, ²³²:

1. There is concordance between animal studies and human epidemiological observations that PFAS modify the immune response, and
2. There are noted complexities in assuming dose-response continuums, including possible differences in life-stage vulnerability.
3. Authors of these reviews note uncertainty about which outcome will be of most importance but agree that immunotoxicity should be included among sensitive human PFAS toxicity endpoints.

PFAS and Cancer

Numerous studies have investigated PFAS carcinogenicity, mainly focusing on perfluorooctanesulfonic acid (PFOS) and perfluorooctanoic acid (PFOA). Perfluorohexanoic acid (PFHxA) is the only other PFAS that has been investigated in an animal study and null findings were reported ²³³. Human studies for PFOS and PFOA include chemical workers, communities with contaminated drinking water, and the general population. A 3.3-fold increase in prostate cancer mortality was reported for each month spent in the chemical division with PFOA production was observed among occupationally exposed workers, but the number of cases was small ²³⁴. Later data from this occupational cohort did not support an association between occupational exposure and cancer mortality or incidence ²³⁵. The strongest evidence for increased cancer risk has been reported by studies among community members whose drinking water was contaminated by PFOA. Barry et al ²⁰⁰ and Vieira et al ²⁰¹ showed a positive association between PFOA levels and kidney and testicular cancers among participants in the C8 Health Project. These studies form the foundation of the overall conclusion from the C8 Health Project. Results among studies conducted in general population are inconsistent. Eriksen et al ²³⁶ was a the first to examine PFOA exposure and cancer in the general population and they did not find an association between plasma PFOA or PFOS concentration and prostate, bladder, pancreatic or liver cancer.

The International Agency for Research on Cancer (IARC) concluded that perfluorooctanoic acid (PFOA) is possibly carcinogenic to humans (Group 2B) and the United States Environmental Protection Agency indicated that there was suggestive evidence of cancer causing potential of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) in humans ¹⁰⁵, ¹⁰⁶, ¹⁰⁷, ¹⁰⁸. Increases in testicular and kidney cancer were noted in highly exposed humans ¹⁰⁹, ¹¹⁰, ¹¹¹.

PFAS in Kidney disease, uric acid, and kidney cancer

Extended human half-lives of long-chain PFAS are attributed to active renal tubular reabsorption. Legacy PFAS such as PFOA and PFOS are concentrated in renal tissues, and histopathologic, molecular, oxidative stress, and epigenetic studies provide evidence of potential nephrotoxicity ²³⁷, ²³⁸, ²³⁹, ²⁴⁰. In addition, the strong influence of kidney reabsorption on the extended half-lives of long-chain PFAS is consistent with both human protein binding and experimental PFAS excretion data.

Human studies have associated legacy PFAS exposure to diminished glomerular filtration and/or defined chronic kidney disease in adults and children ²⁴¹, ²⁴², ²⁴³, ²⁴⁴. However, this outcome may be due to reverse causation ²⁴⁵, ²⁴². Some reviews of the available epidemiologic and toxicologic evidence suggest causative links between PFAS and diminished kidney function and chronic kidney disease ²³⁹, ²⁴⁶; these authors also note several knowledge gaps and uncertainty about which proposed mechanisms of action are most important. A propensity score approach to NHANES data ²⁴⁷, ²⁴⁸ and a study with repeated PFAS and health measures over an 18-yr period ²⁴⁴ recently concluded that PFAS exposure likely causes diminished renal glomerular filtration.

Uric acid, a biomarker of increased risk for renal disease ²⁴⁹, is also consistently associated with PFAS exposure in adults and children ²⁵⁰, ¹¹⁸, ¹⁸³, ²⁴³, ²⁵¹, ²⁵², including a visible dose-response curve that begins at or near historic background levels in human populations ²⁵⁰, ²⁵². Serum PFAS concentrations exhibit an inverted U-shaped pattern related to glomerular filtration, initially exhibiting a modest accumulation as glomerular filtration begins to decrease and then decreasing in advancing renal disease, likely due to failure of normal strong reabsorption mechanisms in moderate to severe kidney disease ²⁴⁷. This finding is more dramatic across stages of glomerular filtration when there is also albuminuria ²⁵³. Studies suggest that the association of PFAS to uric acid is not due to reverse causation and is underestimated because the failing kidney excretes long-chain PFAS but retains uric acid. An implication is that population outcomes that occur in the presence of either albuminuria or moderate to severe renal disease such as hypertension ²⁵⁴ increasing presence of and uric acid (a biomarker of renal disease) ²⁵⁵, ²⁵² can be underestimated in cross-sectional studies; in other words, the link between these health outcomes and PFAS exposure is obscured in these studies because of enhanced PFAS excretion patterns in the presence of either albuminuria or moderate to severe kidney disease. Furthermore, the strong influence of renal reabsorption on the long half-lives of long chain PFAS is consistent with both human protein binding of PFAS and experimental PFAS excretion rates in high-dose rodent studies ²⁵⁶.

Kidney cancer diagnoses have been increasing since 1975, a finding that is partially independent of improved detection, with 5-year cancer-specific survival of approximately 80% ²⁵⁷. The C8 Health studies noted longitudinal (n = 32 254) increases of kidney cancer (hazard ratio = 1.10) and kidney cancer mortality ²⁵⁸, ²⁰⁰, ²⁰¹. A review of 6 published studies found long-chain PFAS exposure associated with kidney cancer or kidney cancer mortality, with risks ranging from 1.07 to 12.8 ²³⁹. Subsequent preliminary data from the heavily exposed Veneto, Italy, population also suggest a significant increase in kidney cancer mortality with PFAS exposure ²⁵⁹. Evidence is accumulating for PFAS as a cause of chronic disease and kidney cancer.

PFAS Thyroid effect

The C8 Science Panel concluded that there is a “probable link” of PFOA exposure to thyroid disease, with sex-specific outcomes in women (for hyperthyroid disease) versus men (hypothyroid disease) ²⁶⁰. Subsequent reviews drew attention to hypothyroid outcomes in women and children and to the possibility that populations with a priori circulating antithyroid peroxidase antibodies may be at additional risk ²⁶¹. A broad childhood disease review noted “some evidence” that PFAS cause childhood hypothyroidism and characterized the number of studies as “limited” for childhood disease conclusions ²⁶². A meta-analysis of 12 child and adult studies that excluded populations with higher exposures noted that PFAS exposure is negatively associated with serum total thyroxine levels and that “PFAS could induce thyroid dysfunction and disease” ²⁶³.

Human thyroid disease is mostly the result of an autoimmune response and is 5 to 10 times more prevalent in women than men ²⁶⁴. Concerning PFAS and clinically diagnosed outcomes, women in the highest quartile of PFOA exposure (>5.7 ng/mL) reported clinical hypothyroid disease (odds ratio 2.2) over 3 cycles of National Health and Nutrition Examination Survey (NHANES) data (1999–2006, n = 3974 adults), with similar findings in men ²⁶⁵. The C8 Science Panel studies (median serum PFOA 26.1 ng/mL) found thyroid disease hazard ratios of 1.00, 1.24, 1.27, 1.36, and 1.37 across cumulative exposure quintiles in women ²⁶⁶, with parallel hypothyroid findings in children aged 1 to 17 years ²⁶⁷. The Ronneby, Sweden, population experienced excess risk of thyroid disease in a discrete time period (1984–2005) among women (hazard ratio 1.29) that did not persist over time despite higher cumulative PFAS exposure ²⁶⁸. The authors did not link exposure to hypothyroid outcome, noting a nonmonotonic dose-response relationship ²⁶⁸.

Human population studies augment experimental data that PFAS interact with thyroid hormone binding proteins, one of several mechanisms by which PFAS can perturb feedback relationships between free thyroid hormone and the hypothalamic-pituitary-thyroid axis ²⁶⁹, ²⁷⁰, ²⁷¹. Exposures to PFAS also interfere with thyroid peroxidase (TPO) enzyme activity in vitro ²⁷². Several PFAS studies have pursued this putative mechanism, finding that maternal and neonatal thyroid hormone outcomes were more readily detected in those with a priori abnormally high circulating anti-TPO antibodies ²⁷³. One case-control study investigated congenital hypothyroidism, a rare condition ²⁷⁴. Serum concentrations of PFOA (5.40 vs 2.12 ng/mL), perfluorononanoic acid (PFNA; 1.93 vs 0.63 ng/mL), perfluorodecanoic acid (PFDA; 0.52 vs 0.30 ng/mL), and perfluoroundecanoic acid (0.98 vs 0.44 ng/mL) were higher in the newborns with congenital hypothyroidism; and levels of several PFAS, including PFOA and PFHxS, were correlated with thyroid autoantibodies ²⁷⁴.

Thyroid disease is not the only concern. Doctors are concerned about subclinically elevated thyroid-stimulating hormone (TSH) in early pregnancy because it may be associated with several possible adverse maternal and fetal outcomes ²⁷⁵. This general concern has prompted numerous PFAS-exposure evaluations of corresponding TSH in maternal serum, cord blood, and newborns. A review of maternal and child biomarkers with PFAS exposure noted that higher TSH has been reported in 4 second-trimester studies ²⁷⁶, but there are also conflicting findings. Studies measuring PFAS in the first trimester have also found associations between PFAS exposure and altered TSH levels in newborns, including nonmonotonic patterns of dose response that mirror the marked alterations of thyroid hormone levels during pregnancy ²⁷⁷.

From the available studies, PFAS definitively alter human thyroid hormones and potentially contribute to thyroid auto-immunity but do not so far appear to be a cause of thyroid cancer ²⁰⁰, ²⁰¹. Also, thyroid cancer is usually survived; thus, morbidity rather than mortality studies are useful.

PFAS Reproductive and Developmental effects

Exposure to PFOA impairs human sperm motility and sperm penetration into viscous media ²⁷⁸, ²⁷⁹ and is longitudinally associated with lower sperm concentration and count and higher adjusted levels of luteinizing and follicle-stimulating hormones in young men ²⁸⁰, ²⁸¹, ²⁸². Serum concentrations of PFAS are also cross-sectionally associated with deleterious markers of semen quality ²⁸³, ²⁸⁴.

Legacy and emerging PFAS have been found in follicular fluid ²⁸⁵. They appear to alter endometrial regulation such as progesterone activity in young women ²⁸⁶ and possibly menstrual cycle length ²⁸⁷. Associations with menarche and menopause may be substantially due to reverse causation because menstruation is a route by which women eliminate PFAS ²⁴⁵, partially explaining why men have higher PFAS levels than women in the same communities. Women on birth control and who do not menstruate or with poor cyclicity because of age, activity level, or disease may have elevated PFAS levels in comparison with menstruating women. Exposure to PFAS has been associated with endometriosis in the United States and in China ²⁸⁸, ²⁸⁹, but the specific PFAS associated with this effect vary among studies.

Time-to-pregnancy (fecundity) studies provide indirect evidence of changes in fertility. Methodologic considerations include maternal and paternal age, parity (which in turn affects serum PFAS), and health status. Among 1240 women in the Danish National Birth Cohort, PFOS exposure was associated with decreased fecundity (median serum PFOS 35.5 ng/mL) ²⁹⁰. Reverse causation may explain this finding because it is duplicated in parous, but not among nonparous, women ²⁹¹, ²⁹². Prospective odds of actual infertility in the Maternal-Infant Research on Environmental Chemicals cohort (n = 1743) at low-dose exposures were associated with PFOA (geometric mean 1.66 ng/mL; odds ratio = 1.31) and PFHxS (odds ratio = 1.27) ²⁹³. The reported fertility rate improved following water filtration in a PFAS-contaminated community along with measures of birth weight ²⁹⁴.

Per- and polyfluoroalkyl substances (PFAS) reliably move across the placenta and enter breast milk ²⁹⁵, ²⁹⁶; serum PFAS levels in young children generally exceed maternal serum concentrations ²⁹⁷, ²⁹⁸, ²⁹⁹. Population studies provide evidence that breastfeeding duration and milk quantity are adversely affected by PFAS exposure ³⁰⁰, ³⁰¹, ³⁰².

A systematic review reported that PFOA exposure was associated with a small decrease in infant birthweight; the meta-analysis estimated that a 1-ng/mL increase in PFOA was associated with an approximately 19-g reduction in birth weight ³⁰³. The authors noted similarities in experimental studies ³⁰⁴, ³⁰⁵ and concluded that there was “sufficient” human and corroborative toxicology evidence of a detrimental effect of PFOA on birth weight ³⁰⁴, ³⁰⁵, ³⁰³. However, another meta-subpopulation analysis, focused on early pregnancy or the time shortly before conception, detected only a small and nonsignificant association, which was less subject to bias ³⁰⁶. Different approaches to the possible confounding role of shifting glomerular filtration rates in pregnancy can affect interpretations; evidence suggests this consideration can, at most, only partially explain associations of PFAS exposure to decreased birth weight ³⁰⁷, ³⁰⁸. A recent review of mostly prospective cohort studies (n = 24 studies) noted PFAS associated with altered fetal and postnatal growth measures, such as lower birth weight. Many (n = 22) of the relevant studies suggest developmental and childhood immunomodulatory effects, whereas 21 studies concerning neurodevelopment were inconclusive ²¹⁰. The authors of the review noted methodologic challenges of developmental and newborn epidemiology, including consideration of critical exposure windows for developmental effects, the effects of breastfeeding and parity on maternal PFAS levels, and the variety of possible mechanistic explanations for growth outcomes, such as disruption of glucocorticoid and thyroid hormone metabolism in utero ²¹⁰. Recent Faroe Island studies report that prenatal PFAS effects on thyroid hormone status do not support a causal relationship ³⁰⁹.

Review articles suggest that prenatal exposure to PFOA may increase risk of subsequent childhood adiposity, noting that steroid hormones, retinoid X receptor, and other pathways may be contributing to this effect ³¹⁰, ³¹¹. Prospective evidence supports this relationship in adults with a high risk of diabetes ¹⁵⁵. However, some well-performed community studies do not support this outcome in adults or children ³¹², ³¹³.

Based on several preliminary findings, supported by longitudinal follow-up studies ³¹⁴, ³¹⁵, ³¹⁶, ³¹⁷, ³¹⁸, the C8 Science Panel concluded that PFOA is probably linked to pregnancy-induced hypertension or preeclampsia. Population-level evidence implicating additional PFAS having pregnancy-induced hypertension (preeclampsia) has included studies with longitudinal designs ³¹⁹, ³²⁰, ³²¹. Experimental support includes PFAS effects on human trophoblast migration in vitro ³²² and recent evidence of PFOA and GenX (or hexafluoropropylene oxide dimer acid) effects on mouse placenta, as well as excessive gestational weight gain ³²³. However, a recent longitudinal study did not find an association of PFAS with pregnancy-associated hypertension (preeclampsia) ³²⁴.

The possibility that circulating PFAS may reduce bone mineral density (BMD) has been investigated. Cross-sectional and practical trial associations have been found in adults ³²⁵, ³²⁶, ³²⁷ and there is emerging longitudinal evidence from a mother and child pair study indicating that children may also be affected ³²⁸.

Testicular cancer diagnoses are increasing steadily, a trend unrelated to improved detection ³²⁹, ³³⁰. Most patients diagnosed (>90%) will be cured and die of other causes; mortality studies therefore provide little help in understanding disease risk factors. The C8 Science Panel detected longitudinal evidence for increased testicular cancer risk (1.35, 95% CI 1.00–1.79) for cumulative PFOA exposure ²⁰⁰. There are ample supportive data of testicular damage following PFAS exposure, including strong evidence of endocrine disruption; but the cell-specific associations are different in humans (germ cell) than the outcomes in rodents (stromal) ².

Per- and polyfluoroalkyl substances have deleterious effects on conception, pregnancy, and infant development ². The underlying birth weight data are mostly supportive, although the subsequent growth and adiposity literature is mixed ². The most sensitive reproductive and developmental outcomes are a topic of ongoing discussion.

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