Perfluorinated compounds (PFCs) have been in use for over 60 years in a wide array of applications. PFCs were first manufactured in the US from about 1947, with perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) as primary products . PFC was later found to contaminate ground and surface water, and PFOS was found to accumulate in freshwater fish . These compounds possess a strong carbon-fluorine bond, which leads to persistence of the PFCs in the environment and the human body . Thus, the high thermal, chemical and biological inertness that make the PFCs useful for many industrial purposes at the same time also generates an environmental hazard.
Serum-PFC analyses conducted by the Centers for Disease Control and Prevention (CDC) show that PFOS and PFOA are detectable in virtually all Americans , with children often showing higher serum concentrations than adults . Analyses of paired samples of maternal serum and cord serum show that PFCs are transferred through the human placenta [5, 6]. Due to global dissemination of PFCs, their serum concentrations in children and pregnant women even in the remote locations, such as the Faroe Islands , are similar to US levels. Exposures to some PFCs in the Faroes may occur primarily through marine diets . Despite the extensive use of these compounds for many decades, and the persistence and cumulative properties of the PFCs, the toxicology data base is still incomplete and has allowed only preliminary risk assessments so far.
Using animal toxicity data, calculations of benchmark dose levels (BMDLs) have been carried out for a 10% deviation relative to control values (i.e., a Benchmark Response or BMR of 10%); they resulted in serum concentrations of 23 mg/L and 35 mg/L for PFOA and PFOS, respectively [9–11]. Toxicokinetic modeling and standard assumptions about water intake then allow derivation of acceptable drinking water levels [11, 12]. So far, the U.S. Environmental Protection Agency (EPA) has issued a draft risk assessment of PFOA in 2005, but no final version has yet been published, nor has a Reference Dose (RfD) been defined. However, the EPA has issued provisional health advisories of 0.4 μg/L (400 ng/L) for PFOA and 0.2 μg/L (200 ng/L) for PFOS in drinking water . Similarly, the Agency for Toxic Substances and Disease Registry concluded in its draft toxicological profile in 2009 that there was insufficient evidence at the time to develop a minimal risk level . For chronic exposure, state authorities have issued limits for PFC concentrations in drinking the water, e.g., in Minnesota , where the limit for both PFOS and PFOA is 0.3 μg/L (300 ng/L). The limits were based on PFOS effects on the liver and thyroid, and PFOA effects on the liver, fetal development, reduction in red blood cell numbers, and immune system changes in experimental studies . A lower guidance limit of 0.04 μg/L (40 ng/L) has been determined for PFOA by the state of New Jersey . Other agencies, such as the European Food Safety Authority  have recommended similar exposure limits that relied on the same toxicology data while using different default assumptions.
PFC toxicity in animal models at first suggested the liver as a main target organ, but so far chronic toxicity data only in the rat have been published [1, 12, 17, 18]. However, recent evidence suggests that toxicology outcomes used in derivation of exposure limits may not represent the most sensitive endpoints. Thus, interference with mammary gland development in mice with developmental exposure seems to occur at low exposures; benchmark dose calculations using a variety of models showed that a 10% BMR corresponded to a serum-based BMDL for PFOA of 23–25 μg/L (or ng/mL) [12, 17]. This BMDL differs by a factor of 1,000 from the previously mentioned BMDL based on liver toxicity (i.e., 23 mg/L or 23,000 μg/L). Thus, current limits for PFOA in drinking water based on the latter value may not be as protective as intended, despite the use of uncertainty factors.
Likewise, immunotoxicity of PFCs has been demonstrated in rodent models, avian models, reptilian models, and mammalian and nonmammalian wildlife . For example, in a commonly used mouse model, PFOA effects include decreased spleen and thymus weights, decreased thymocyte and splenocyte counts, decreased immunoglobulin response, and changes in specific populations of lymphocytes in the spleen and thymus. Reduced survival after influenza infection has also been reported as an apparent effect of PFOS exposure in mice . Another study found that the lowest observed effect level (LOEL) for males corresponded to an average serum-PFOS concentration of 92 ng/g (about 94 μg/L), though 7-fold higher in females . The LOEL serum concentration in males is similar to typical levels found in serum samples from subjects exposed to contaminated drinking water .
Given the concern about immunotoxicity as a possible critical effect  and the possibility of developmental toxicity , studies in child populations have recently focused on antibody responses to childhood immunizations as a clinically relevant parameter that reflects major immune system functions . The subjects have all received the same doses of vaccine antigens at the same ages and can then be examined at similar ages, i.e., similar intervals after the most recent vaccination . Our studies focused on the fishing community of the Faroe Islands , and these prospective population data  seem appropriate for calculating benchmark doses as a contribution to future risk assessments.
While benchmark dose calculations from toxicology data are fairly straightforward, using epidemiological studies can be more complicated due to the need for covariate adjustments . In addition, decisions on dose–response models may be crucial, as a null exposure group is usually not available, thus requiring extrapolations beyond the exposure interval observed.