Making the Connection Between PFASs and Agriculture Using the Example of Minnesota, USA: A Review

Abstract

Exposure to per- and polyfluoroalkyl substances (PFASs) can cause detrimental health effects. The consumption of contaminated food is viewed as a major exposure pathway for humans, but the relationship between agriculture and PFASs has not been investigated thoroughly, and it is becoming a pressing issue since health advisories are continuously being reassessed. This semi-systematic literature review connects the release, environmental fate, and agriculture uptake of PFASs to enhance comprehension and identify knowledge gaps which limit accurate risk assessment. It focuses on the heavily agricultural state of Minnesota, USA, which is representative of the large Midwestern US Corn Belt in terms of agricultural activities, because PFASs have been monitored in Minnesota since the beginning of the 21st century. PFAS contamination is a complex issue due to the over 14,000 individual PFAS compounds which have unique chemical properties that interact differently with air, water, soil, and biological systems. Moreover, the lack of field studies and monitoring of agricultural sites makes accurate risk assessments challenging. Researchers, policymakers, and farmers must work closely together to reduce the risk of PFAS exposure as the understanding of their potential health effects increases and legacy PFASs are displaced with shorter fluorinated replacements.

1. Introduction

PFASs are defined as aliphatic substances which contain at least one carbon atom on which all hydrogen atoms have been replaced with fluorine atoms. Perfluoroalkyl sub-stances are substances in which all hydrogen atoms attached to the carbon chain have been substituted by fluorine, and polyfluoroalkyl substances are not fully fluorinated [1]. The Organisation for Economic Co-operation and Development proposes a broader definition by suggesting that any compound containing at least one fully fluorinated methyl or methylene group be referred to as a PFAS [2]. There are currently over 14,000 chemicals listed as PFASs in the United States Environmental Protection Agency toxicity databases [3]. Some of the most relevant representatives are presented in

Table 1. A list of the most relevant PFASs. * The ammonium salt of HFPO-DA is also known as GenX.

Name	Abbreviation	Chemical Formula	Structural Formulas
Perfluorobutanoic acid	PFBA	C4HF7O2
Perfluorobutanesulfonic acid	PFBS	C4HF9O3S
Perfluorohexanoic acid	PFHxA	C6HF11O2
Perfluorohexanesulfonic acid	PFHxS	C6HF13O3S
Perfluorooctanoic acid	PFOA	C8HF15O2
Perfluorooctanesulfonic acid	PFOS	C8HF17O3S
6:2-Fluorotelomersulfonic acid	6:2 FTS	C8H5F13O3S
Hexafluoropropylene oxide dimer acid *	HFPO-DA *	C6HF11O3
6:2 chlorinated polyfluoroalkyl ether sulfonate	F-53B	C8ClF16KO4S

.

Perfluoroalkyl and polyfluoroalkyl substances (PFASs) are ubiquitous and can be detected in the environment, wildlife, human serum, and tissue [4,5,6,7]. PFASs can accumulate in the human body over time and cause different toxicological effects [8,9], such as cancer, kidney and liver disease, and adverse effects on reproductivity and development [10]. The mean half lives of some of the most common PFASs, namely perfluorooctane sulfonic acid (PFOS), perfluorohexane sulfonic acid (PFHxS), and perfluorooctanoic acid (PFOA), are 5.4 years, 8.5 years, and 3.8 years, respectively [11].

PFASs can be found in many products, such as firefighting foams, food packaging materials, household products, medical devices, pesticide formulations, and surfactants [12]. The widespread industrial and commercial utilization of these substances is based on their chemical and physicochemical properties. The carbon–fluorine bond is the strongest bond in organic chemistry, with a dissociation energy of up to 536 kJ/mol [13]. Therefore, the PFAS family of chemicals shows exceptionally high chemical and thermal stability [14,15,16]. Due to their polar nature, most PFASs show high solubility in water [17] and are therefore environmentally mobile [18]. Four main activities are responsible for the contamination of PFASs in the air, water, and soil. These include leakages and emissions of PFASs at manufacturing sites, use of aqueous film-forming foam (AFFF) for firefighting activities, leaching from landfills, and the application of biosolids for wastewater treatment [19,20,21,22].

Although PFASs have multiple potential exposure paths to humans, a significant exposure route of concern is the consumption of contaminated food [23]. PFASs can accumulate from the environment in crops and produce that are consumed directly by humans [24], and the PFASs in fodder and grain used to feed livestock can lead to the contamination of the animals and their related food products [25,26]. In Maine, USA, dairy farms were shut down due to PFAS contamination in their products, with PFASs being detected on more than 50 farms [27,28,29,30]. Most environmental research has primarily focused on PFAS sources and the fate of these substances in the air, water, and soil [31]. Research on downstream biological systems, such as agricultural crops and livestock, experiencing PFAS contamination is lacking. Understanding the relationship between PFASs in the environment and agriculture is critical as humans are dependent on food production and approximately 44% of habitable land surface is used for agricultural purposes [32].

The research objective of this study is to detail the sources of PFAS contaminations in agriculture, determine the environmental fate of PFASs, and outline the connection to agriculture to fill knowledge gaps. The history, contamination levels, and actions regarding PFASs are discussed for a typical Midwestern state (Minnesota, USA) so that the findings can be put into the historical context, and other agricultural areas can assess their paths towards protecting their agricultural sector from contamination issues. This work aims to make meaningful contributions to public health by reducing the risk of PFAS exposure by suggesting sensible actions based on our findings. Furthermore, we suggest future research directions to close dangerous knowledge gaps.

2. Materials and Methods

A semi-systematic literature was carried out to address the link between PFAS contamination and agriculture. Different databases and search engines (i.e., SciFinder (web-based, Chemical Abstracts Services, Columbus, OH, USA), Scopus (web-based, Elsevier, Amsterdam, Netherlands), Web of Science (web-based, Clarivate, Philadelphia, PA, USA), PubMed (web-based, U.S. National Library of Medicine, Bethesda, MD, USA), Google Scholar (web-based, Google LLC, Mountain View, CA, USA), and ScienceDirect (web-based, Elsevier, Amsterdam, Netherlands)) were utilized to identify relevant peer-reviewed articles. Governmental databases (e.g., Minnesota Groundwater Atlas) and reports (e.g., Minnesota’s PFAS Blueprint) were examined from multiple agencies to gather Minnesota-related and other important information. Various single and combined search terms were employed due to the complexity and diversity of the presented research objective. For example, the number of publications per year for a typical query (i.e., PFAS * and agriculture *) is shown in Figure 1. Excluding most sources from Section 3.1.1, the majority of relevant papers were published from 2019 onward and no publications were found prior to 2007 regarding PFAS in agriculture specifically. The exponential growth of the number of publications indicates the increasing scholarly attention regarding this issue to address pressing knowledge gaps and lacking mitigation strategies.

Multiple qualitative criteria were used for the inclusion or exclusion of the respective sources. These encompassed the (1) relevance, (2) originality (primary and field studies were preferred), (3) publication date (papers published between 2000 and 2024 were considered; this does not apply to historic documents and a few exceptions), (4) overall quality (e.g., methodology, sample size, transparency, etc.), and (5) impact (i.e., number of citations or legislative documents).

3. Results and Discussions

3.1. PFASs in Minnesota

3.1.1. History of PFASs in Minnesota

Minnesota’s experience with PFASs is deeply intertwined with the beginning of PFAS production in the 1940s. PFOS- and PFOA-related substances and other fluorochemicals were manufactured at Minnesota Mining and Manufacturing (officially renamed 3M in 2002) production facility in Cottage Grove, MN [33]. As early as 1955, research indicated that PFOS could impact biological systems, as protein binding properties were observed in PFOS [34]. In 1961, DuPont researchers found that these chemicals can potentially increase the liver size in rats and rabbits [35]. In 1997, 3M was selling almost 1900 metric tons of PFASs in the US. [36]. Waste products, sludges, and wastewater were generated, disposed of at landfills or discharged to the Mississippi River [37]. In 2002, 3M informed the Minnesota Pollution Control Agency (MPCA) that PFOA and PFOS had been detected in groundwater reservoirs used near production and disposal sites [38].

A number of locations linked to PFAS wastes in and around Cottage Grove, MN, were investigated to examine the extent of contamination [39]. While Cottage Grove and the surrounding areas with contamination were more rural areas when 3M began disposal of PFAS wastes, their population grew to become part of the greater Twin Cities (Minneapolis–St. Paul) metropolitan area. According to the MPCA, 3M’s manufacturing activities led to the contamination of groundwater covering more than 150 square miles, which affected the drinking water quality of more than 140,000 Minnesotans. Different PFASs were detected at varying concentrations in all samples, including groundwater, surface water, municipal wells, sediment, sludges, influent/effluent, fish, plants, leachates, and gas condensates at landfills [37,39]. Further research analyzed the PFAS contamination in the air [40], fish and bald eagles [41,42], leachate and gas condensates from landfills [43], and yard waste sites [44]. Statistical evaluations showed that residents of affected communities (e.g., Oakdale) were 30% more likely to be diagnosed with prostate cancer, and infants were 34% more likely to be born with low birth weight compared to unaffected communities [45].

3M decided to terminate the manufacturing of PFOS- and PFOA-related compounds after the initial discovery of contamination [5,46]. In 2010, the Minnesota Attorney General sued over contaminated drinking water and natural resource findings [47]. More recently, 3M announced that it will cease all PFAS manufacturing and stop selling products containing PFASs by 2025 due to the “rapidly evolving external regulatory […] landscape” [48]. Communities are still tremendously concerned about PFAS exposure and the associated health effects and urgently demand that appropriate actions be taken.

3.1.2. Contamination in Minnesota

The detection of major contamination around the Twin Cities metropolitan area led to a larger effort to examine PFAS contamination and spread than is seen in other communities and states. While Cottage Grove was the single most sampled area, rural areas throughout Minnesota were also heavily sampled. Much of the data has been reported to the MPCA, who publishes it as part of their role in investigating and regulating pollution in the state [49]. With this extensive data set, Minnesota provides a temporally and geographically complete example for assessing issues related to PFAS contamination. That said, these datasets are complex and difficult to interpret because of the many unique PFAS chemical species and behaviors, in addition to changing testing protocols as the field has advanced.

Drinking Water Standards as a Guide

The potential human health risks of consuming contaminated drinking water have meant that most Minnesota and nationwide PFAS investigations have concentrated on the contaminant levels in drinking water [49,50]. However, the extent of the problem has meant that Minnesota has expanded testing to surface and groundwater statewide. While the dangers represented by contaminated drinking water are different than agricultural PFAS contamination, water data provides the best guide to understanding contamination in agriculture. The vast PFAS water data from different sites and examining different types of PFASs require that a metric be established to help evaluate locations where contamination levels are potentially detrimental to health.

The Health Index (HI) was adopted by Minnesota in 2002 to evaluate the associated risk of consuming drinking water, and it is continuing to be updated as new data emerges [51]. The HI uses drinking water guidance values for perfluorobutanoic acid (PFBA) (7.000 µg/L), perfluorobutanesulfonic acid (PFBS) (2.000 µg/L), perfluorohexanoic acid (PFHxA) (0.200 µg/L), PFHxS (0.047 µg/L), PFOA (0.035 µg/L), and PFOS (0.015 µg/L) (see Equation (1)) [52].

HI = ([PFBA])⁄(7 µg/L) + ([PFBS])⁄(2 µg/L) + ([PFHxA])⁄(0.2 µg/L)
+ ([PFHxS])⁄(0.047 µg/L) + ([PFOA])⁄(0.035 µg/L) + ([PFOS])⁄(0.015 µg/L)

(1)

If the HI is greater than 1, the drinking water is considered to have potentially adverse health implications. More information on the regulation of PFAS contamination levels is provided in Section 3.5.

As of July 2024, 919 Community Water Systems out of 966 have been tested by the MPCA, with 5 exceeding the proposed HI. The contaminated Community Water Systems are in larger urban areas, except for Swanville, a small community in central Minnesota. Most Community Water Systems showing PFAS contamination are located in the Twin Cities metropolitan area, which can be linked to 3M production in Cottage Grove (see Section 3.1.1) [49].

The geospatial distribution of wells that contain at least one PFAS that exceeds health-based guidance values and their proximity to wellhead protection areas, which surround public water supply wells and where groundwater contributes to the respective wells, are shown in Figure 2. Out of 13,884 locations/timepoint sampling combinations, 17% contained at least one exceedance and 4189 total exceedances were detected [53]. These were found both in urban and rural/agricultural regions (see Figure 2).

Air Pollution

The first extensive study on PFAS levels in ambient air was published by the MPCA in April of 2022 [40]. Researchers sampled the ambient air at four different sites. Three sites were located near PFAS-emitting sites in urban areas (Duluth, St. Louis Park, and Eagan), while the fourth site (Grand Portage) was in a rural area and was expected to be a low-PFAS reference site. During the observation duration, 17 different PFASs were detected, varying in concentration and time of observation. The MPCA found PFBA, PFBS, PFOA, and PFOS in all air samples, with PFBA contributing approximately between 47% and 70% of the total PFAS concentration. The authors attribute the abundant occurrence of PFBS to the direct emission and degradation of precursors in the atmosphere. Precursors are used as intermediates in the manufacturing process of surfactants or fluoropolymers. They are also used as substitutes for chlorofluorocarbons, coolants, solvents, or fire suppressors [54]. Additionally, 6:2 fluorotelomer sulfonate is known to replace PFOS at chrome plating facilities (e.g., at the St. Louis Park site) and is additionally applied in AFFFs. Therefore, 6:2 fluorotelomer sulfonate was regularly detected in ambient air [40].

Though Grand Portage was chosen due to its remote location, low population (684 in 2019), and limited industry, it showed the second highest airborne PFAS contamination with a mean concentration of about 100 pg/m³. The authors hypothesize that this surprising finding is likely due to emissions from the local fire department and the solid waste transfer station. They also speculate that long-distance atmospheric transport might contribute to the relatively high concentrations. The findings that the rural community of Grand Portage could have high detectable levels of airborne PFAS indicates that rural/agricultural areas should not assume that their relative isolation is potentially sufficient to limit exposure to airborne PFASs.

3.2. Significant Sources of PFASs in the Environment

3.2.1. Release of PFASs During Their Production, Use, and Disposal Phases

Manufacturing plants that produce PFASs or use PFASs in production of other products contribute significantly to environmental contamination by PFASs [55]. During the manufacturing process, PFASs can be emitted into the atmosphere [40,56] or discharged as contaminated wastewater that is often not sufficiently treated at wastewater treatment plants before it enters the aquatic environment [37,55]. Hazardous wastes containing PFASs can also be generated during the production process, which may not always be mitigated when incinerated or disposed of; improper storage or burial might contaminate the environment [57].

PFASs can enter the environment during the use of the PFAS-containing products [15]. For example, some carpets contain PFASs to repel stains, which might be released from carpet fibers over time [58]. Functional textiles (e.g., outdoor clothing) are also treated to have water-repelling properties. They are exposed to radiation, rain, heat, and abrasive stress that can cause the emission of repellents [59]. Many non-stick pans are coated using PFAS precursors, which can flake off during cooking or during the cleaning process. PFASs are present in more than 200 use categories [15], which have the potential to lead to unintentional PFAS release.

After their use, roughly 146 million metric tons (50%) of solid wastes were landfilled in 2018 in the US. [60]. Solid waste management facilities are viewed as important point sources of PFAS pollution, and thus, many researchers have investigated the PFAS concentration in landfill leachates [61,62,63]. PFASs undergo long-term leaching, and precursors degrade [64] as they are exposed to sunlight, microbial activity, varying redox conditions (available oxygen, pH-value, content of redox active species, etc.), or wastewater treatment [65]. PFAS leachate and gas condensates at solid waste landfills were evaluated across the state by the MPCA between 2005 and 2008. Down-gradient (points where water flows away from landfill) and upgradient (points where water flows to the landfill) groundwater samples were taken and found to have significant differences in concentrations of PFBA and PFOA, with PFAS concentrations being higher in the water leaving the sites. PFOS was detected in 10% of the samples. There were noteworthy differences between the median PFOA and PFBA concentrations in the groundwater at lined and unlined landfills. The median concentrations at lined landfills were 2.39 ng/L and 33.2 ng/L, respectively. The median concentrations of PFOA and PFBA at unlined landfills were 0.36 ng/L and 7.29 ng/L, respectively. The same study also showed that gas condensates were contaminated [43].

In total, the MPCA investigated the groundwater at 102 out of 111 landfills of interest. Out of those, 98% showed elevated PFAS levels, and the drinking water guidance values of the Minnesota Department of Health (MDH) were exceeded at 62 sites [66]. The landfills included both metro and rural landfills. Rural landfills are often outside cities and adjacent to active farming operations. Most of the rural landfills analyzed in these efforts are landfills that were closed or consolidated as more stringent environmental rules increased requirements for siting and management. However, these closed landfills are likely to have been leaching PFASs for decades. They are typical of small town/county size landfills found throughout the US Midwest.

Mass balancing PFAS leachates and estimating total releases is challenging as they depend on various fluctuating factors (e.g., rainfall, number/concentration of PFASs) [67]. Lang et al. estimated that the total PFAS mass release from leachates at landfills in the US is approximately 600 kg/year [61].

3.2.2. Aqueous Film-Forming Foam (AFFF)

Over 240 individual PFAS compounds (e.g., perfluoroalkyl carboxylic acids (PFCAs), perfluorosulfonic acids (PFSAs), and perfluoroalkyl sulfonamido compounds) are used in aqueous film-forming foam (AFFF) formulations [68,69] to enhance wetting at the hydrocarbon–water and air–water interfaces [15,69]. In practical terms, this allows firefighting foams with AFFF to more easily encapsulate flammable liquid fuels and starve them of oxygen when attempting to stop a fuel fire. Firefighting and training activities (e.g., at airports, fire training areas, or military sites) have resulted in the contamination of surrounding surface waters and groundwater. However, data describing the lower-volume AFFF use at smaller airports has not been collected, so it is difficult to assess how significant this problem may be at rural airports. As many precursors present in these foams undergo transformation processes [70,71], the fate of AFFF-contaminated sites is not fully understood [68].

Many rural airports are located immediately adjacent to agricultural fields, which provide the open land needed for runway safety. The use of AFFF at these sites, either for training or to combat an actual fire, has the potential to contaminate nearby groundwater and fields.

3.2.3. Biosolids from Wastewater Treatment

Biosolids consist of organic matter that is recovered from sewage during wastewater treatment. They are often applied in agriculture to improve soil fertility by enriching it with nutrients and organic matter [72]. Biosolids are heterogeneous and have varying concentrations of organic matter, microorganisms, bacterial constituents, inorganic materials, and water [73]. As a result, PFASs can bind to various sites of these components with different strengths [74,75,76]. PFAS-contaminated biosolids are generated in the wastewater treatment process because PFASs cannot be efficiently removed from sewage being treated during biosolid separation. Some studies even showed that the concentration of stable perfluoroalkyl acids increases during treatment due to the oxidation of precursors (e.g., fluorotelomer alcohols) [77,78]. Hence, wastewater treatment plants are considered an important point source of PFAS [16]. Although sewage and treated biosolids from residential sanitation systems are known to contain somewhat higher levels of background PFASs, sewage from industrial wastewater treatment is a primary source of more highly contaminated biosolids [79].

Once the contaminated biosolids are applied on a field, the fate of the PFASs is influenced by physicochemical properties of the respective substances and soil. It was shown that the half-lives of different PFASs in soils can range from a few days to years depending on their chain length, functional groups, and their dissociation constant [80,81]. Furthermore, substances can be accumulated by various plants [82,83,84] or leach into the groundwater [23,85]. As a result, PFASs were detected in livestock, and thus, food pro-duction can be an exposure pathway to consumers [86]. Lindstrom et al. investigated well water and surface water close to fields where contaminated biosolids were applied for 12 years and observed elevated PFAS concentrations (up to 11 µg/L for PFOA in surface water) [87].

The Environmental Protection Agency (EPA) estimated that 1.15 million dry metric tons of biosolids were applied to agricultural land in 2021 [88]. The yearly loading of PFASs on agricultural land in the US is estimated at 1375–2070 kg/year [89]. Thus, concerns arise about whether and to what extent biosolids pose a threat to food supply and, therefore, the economic security of farmers [90]. While a typical application of biosolids is not expected to dramatically increase PFAS concentrations in the soil, there is a concern that multiple biosolid applications or the use of heavily contaminated biosolids may increase soil contamination to the point where food or livestock feed would have high levels of PFASs.

Regulatory agencies became aware of the contamination potential of biosolids in 2016, when elevated PFAS levels were detected in milk products of a dairy farm in Maine. The Maine Department of Environmental Protection found combined PFOA and PFOS concentrations of up to 1.420 μg/L. In addition, a PFOS concentration of 32.200 μg/L was detected in dairy products on another farm in 2020 [90].

3.2.4. PFASs in Pesticide Formulations

Approximately 68,000 metric tons of pesticides are applied yearly in the Midwestern US [91]. PFASs have been part of some pesticide formulations as active ingredients or identified as inert ingredients in the past [92]. However, it is not fully clear how many currently used pesticides in the US contain PFASs as active ingredients or for other reasons. Lasee et al. found that 6 out of 10 insecticide formulations contained PFAS concentrations between 3.92 mg/kg and 19.2 mg/kg [93]. However, these results could not be verified in a separate investigation, and no PFASs were detected above the detection limit of 0.2 μg/L [94].

International pesticide active ingredient and application differences also increase the difficulty of quantifying PFAS applications to land. N-ethyl perfluorooctane sulfonamide, also called sulfuramid, is an effective insecticide used to control leaf cutting ants (Atta and Acromyrmex) [95]. It can be transformed into PFOS in the environment. Even though sulfuramid was internationally phased out by 2016 [96], it is still used in South America. Estimates are that Brazilian sulfuramid use has been responsible for the release of up to 487 metric tons of PFOS/perfluorooctanesulfonamide between 2004 and 2015 [97].

Storage containers can also be a PFAS source in pesticides via leaching. Nguyen observed gradual leaching from fluorinated high-density polyethylene containers. The total PFAS concentration was 15 µg/L when methanol was used as a solvent and 3 µg/L in water [98]. However, the relatively small contribution of PFAS from a container to any applied chemical would be difficult to detect. Consider an example of a container that holds 9 L of pesticide covering 4 hectares of crops, which could leach 15 μg PFAS per liter of pesticide, and leaching would result in 34 μg/ha or 3.3 ng/m². This is below the background levels, as noted by Qian and French [94]. Generally, pesticide use is very crop, weather, and geographically specific. Thus, the load of pesticide-related PFASs must be calculated on a case-by-case basis.

3.2.5. PFASs in Rural Versus Urban Communities

It is evident that rural and urban areas are impacted by different PFAS contaminant sources as they vastly differ in their industrial activities, population density, and land use [99]. This is reflected by site-specific contamination patterns, which reveal information about PFAS discharge in the area [100].

The background levels should be lower in rural areas due to reduced human activities, but similarly to urban areas, point sources are primarily responsible for pollution in rural areas. It is important to explicitly discuss sites likely to face contamination issues in agricultural and rural areas. While a significant amount of contamination has been observed in areas near PFAS production sites and in industries utilizing PFASs, such occurrences are relatively uncommon in rural areas. Very few facilities produce PFASs nation-wide, and facilities using large volumes of PFASs in industrial application tend to be closer to urban centers that can supply labor. Therefore, most contamination in rural areas occurs from the application of high-PFAS-concentration biosolids, the use of AFFF, the application of PFAS containing pesticides, or leaching from landfills. While some species of PFASs are spread by air in rural environments, it should not be significantly different than the global background levels. Lastly, in total, there are 321 airports, seaplane bases, and heliports statewide in Minnesota that are generally linked to enhanced contamination levels [101].

3.3. The Behavior and Transport of PFASs in the Agricultural Environment

The fate of PFASs in the environment and its entry into agricultural ecosystems depends on various parameters. The physicochemical properties of particular PFAS species determine their movement, partitioning, and transformation behavior, also called environmental fate [102]. The PFAS chain length [103] and functional groups [104] significantly impact their solubility, mobility, and bioavailability. Thus, each substance has a unique environmental fate, which makes it challenging to develop predictive models in agricultural soils, plants, and livestock [105].

3.3.1. PFASs in the Atmosphere

Non-ionic PFASs (typically precursors) are often sufficiently volatile such that they are transported in the atmosphere [106]. Long-range transport of PFASs, on the order of days to weeks, can occur in the atmosphere. For example, researchers attributed the contamination of the European Arctic to long-atmospheric transport processes, the resulting degradation processes and local sources (e.g., consumer products and AFFFs) [107,108]. The total historical global emissions of PFCAs due to indirect air emissions are estimated to be as high as 350 metric tons, which contributes between 4.8% and 10.9% of the total historical PFCA emissions [109].

Airborne PFASs can migrate to the soil or into surface water via dry (direct deposition) or wet deposition (ab- or adsorption by/at water) [107,110,111,112]. Compared with other types of movement of PFASs, airborne movement likely spreads material over a wider area at a lower concentration. This compares to PFAS movement in water, biosolids, or other solid/liquid mediums. This is likely a primary mechanism for the background PFASs found in many isolated rural communities, both agricultural and non-agricultural.

3.3.2. PFASs in Surface Water, Sediment, and Groundwater

About 40 different PFASs have been detected in the aquatic environment [106]. PFASs with low pKa values and vapor pressure (e.g., PFCAs and PFSAs) are ionic at common environmental pH values. These ionic PFASs are mainly observed in water or bound to particles and sediment [109,113]. The PFASs that persist in the aquatic environment can be transported downstream by rivers until they end up in the marine environment, where PFASs can reenter the atmosphere via sea spray [114]. During this process, the total PFAS concentration decreases downstream until it reaches the oceans [106].

Multiple studies have investigated the sorption behavior of PFASs between water and different sediments/soils [104,113,115,116]. Sediments can store organic pollutants over a long period since sediments show hydrophobic behavior in the aquatic atmosphere [117], and trends of released PFASs (e.g., replacing PFOA with HFPO-DA) can be observed in sediment cores [118]. Other important parameters for sorption are the pH value, organic carbon content, and the coexistence of certain metal cations (e.g., Ca²⁺) [104,113,115,116].

Generally, PFCAs with a chain length shorter than seven carbon atoms were exclusively found in the liquid phase [104], and every extra CF2 moiety on the molecule in-creases binding affinity to the solid phase. Moreover, the sulfonate group enhances the absorption strength further in comparison to the PFCA analogs [116]. Linear PFAS isomers tend to sorb to sediments due to their lower hydrophilicity compared to their branched isomers [119].

Many worldwide studies have investigated groundwater contamination levels [120,121,122,123,124]. Groundwater contamination can be attenuated by the vadose zone due to sorption phenomena [85], and surface water was identified as a key source of groundwater contamination [125,126,127]. In contrast to the atmosphere and deposition on surface water, which retain PFASs for short periods, groundwater (and soil column) contamination reflects the long-term retention of PFASs [128].

The biotransformation of precursors is another important factor in evaluating whether aquifer contamination is a potential source of danger for the environment [121] or human health [9,129]. Biotransformations are highly dependent on the toxic conditions of the soil [70,130,131].

As groundwater enters springs, lakes, rivers, and wetlands or is used for irrigation [132], PFASs that are present can reenter the water cycle [121] and thus might be reintegrated into the biosphere. Overall, PFAS groundwater contamination processes are not well understood, and further research is needed to improve modeling [133].

3.3.3. Behavior and Uptake of PFASs in Soils

PFAS Sorption

The modeling of soil PFAS adsorption is complicated by the wide variation in soil components and the resulting soil’s physical and chemical properties. Soils are a mix of mineral, organic, and biologically active materials and other compounds that can change over the distance of a few meters. These variations interact in concert with the chemical property differences in PFAS species. Some soils simultaneously have surface sites that are charged positively and negatively. The surface charge strongly depends on the pH and on the soil type [134,135]. At the same time, several PFASs occur in an ionic or zwitterionic state (equal number of positively and negatively charged functional groups present in the molecule) in the environment [136,137].

Several studies investigated the vertical distribution of PFASs in the ground after contamination due to biosolids application or AFFF use [6,21,138,139,140]. In most cases, the observation was that the vertical loading of PFASs in the vadose zone decreased sharply with increasing depths. In contrast, Nickerson et al. found that total PFAS concentrations increased along the vertical profile of the soil, and the maximum concentrations were observed at a depth of 3–5 m [141]. The retention and lifetime of PFASs, especially precursors, differ vastly depending on the soil [142,143,144].

The factors influencing the adsorption of PFASs (mainly PFOA and PFOS) onto soils have been thoroughly investigated. It is apparent that the sorption PFASs is not dependent on one factor and that models must use multiple parameters to describe sorption accurately. Organic carbon is often attributed as the driving parameter of PFAS adsorption in the soil [116,145,146,147,148,149].

Another important factor is the content of dissolved inorganic salts in the soil column since they can change the properties of the aqueous phase. Investigations showed a positive correlation between the ionic strength and the sorption of anionic PFAS onto soils [116,146,147,150,151]. It was found that with increasing ion concentrations, the mobility of PFASs along the soil profile decreased (especially for long-chain PFASs with C >10) [151].

Most sorption processes of PFASs might be explainable when considering the organic carbon, pH, and the clay content due to the importance of electrostatic interaction for the sorption processes. More data is needed to conduct multivariate data analysis techniques, which might provide a more accurate understanding [76]. This would be very helpful as values of the field- and laboratory-derived partitioning coefficients for PFOS, PFOA, PFNA, and PFDA in the literature vary by factors of 1.6, 10.9, 1.9, and 1.7, respectively. All field-derived values are larger than the laboratory-derived values [76]. More soil adsorption research is needed, which should include a precise characterization of the soil properties to facilitate future analysis of sorption processes. If the sorption mechanisms are fully understood, farms and areas that are likely to have the potential to have a high concentration of sorbed PFASs can be identified. This ensures effective employment of resources in mitigation efforts.

PFAS Retention

The retention of PFASs in soils depends not only on the interactions between the compound and the soil but also on (1) adsorption at the air–water interfaces, (2) partitioning to soil gases, (3) adsorption at non-aqueous phase liquid (NAPL)–water interfaces, and (4) partitioning to the NAPL [152,153]. The combination of general partitioning mechanisms might not be sufficient to predict retention in some cases [154]. One field study investigated the PFAS partitioning behavior at AFFF-affected sites and observed that at 87% of the examined sites, the concentration levels in the soil exceeded those of the groundwater, indicating the significance of the retention [154].

However, a laboratory study revealed that between 80% and 90% of the ten investigated PFASs were eluted in the flow-through soil column experiments [155]. It is questionable to what extent the results of the study can be transferred to environmental conditions. A further important contributor to PFAS transport is the presence of co-contaminations [156]. Those can increase sorption by providing additional binding sites or lower PFAS sorption by competing for existing sorption sites in the soil [157].

PFASs in Agricultural and Rural Soils

PFAS release mechanisms of concern for agriculture are all point sources of contamination that first impact specific release sites. Thus, preventing agricultural plant and animal systems from being exposed to PFASs requires an understanding of local point sources of contamination and the potential for movement into the agricultural environment.

Depending on factors such as surface and sub-surface water flows and soil types, the areas adjacent to the initial contamination sites can become polluted over time. As described above, the rate of movement is dependent on many different soil, water, and PFAS physical and chemical properties as well as concentration gradients. Based on the data from Minnesota contamination sites, it is likely that sub-surface movement of PFASs from a contamination site can take several years per kilometer of movement in typical soils. This time is reduced in soils with high groundwater flows. In contrast, contamination moves very quickly in surface waters.

The risk of PFAS translocation should be insignificant for the consumer if the products were grown on fields with no history of direct PFAS application and that underwent minimal industrial processing. However, the general lack of monitoring of agricultural sites makes it challenging to perform accurate risk assessments.

3.4. PFASs in Plant and Animal Systems

Agricultural plants and animals are in direct, constant interaction with the outdoor environment, and thus, they can take up PFASs from their surroundings. A better understanding of the uptake mechanism of agricultural plants and animals consuming plants is required to reduce the potential risks of human exposure to PFAS from agricultural products. This section describes the current understanding of these interactions and the potential for contamination of agricultural goods during production.

3.4.1. Plant Uptake of PFASs

PFAS Uptake Mechanisms in Plants

The predominant pathway for plant accumulation is root uptake [158,159,160,161]. Above-ground parts of the plant can also absorb airborne PFASs to a smaller degree [162,163,164]. In the first step of the root uptake process, the contaminants diffuse passively into the surface tissue of the root system. The PFASs are then gradually distributed and accumulate in the root tissue. This only occurs in cell walls that have not sufficiently developed protective layers (e.g., cuticle layer) [165]. If the molecule has sufficient lipid solubility, it can pass the lipid bilayer of the membrane and enter the aqueous phase inside the cell. Charged or polar particles might be hindered from entering the cell membrane by the lipid bilayer, but proteins can function as a transport system [166] due to the high affinity between PFASs and proteins [167]. After they enter the root tissue, PFASs can be absorbed by apoplastic cell wall tissue outside root cell membranes and be transferred through the xylem up to the plant’s shoot (see Figure 3). Here, the Casparian strip inhibits the translocation of long-chain PFCAs, PFSAs, FOSAs, and chlorinated polyfluorinated ether sulfonates by sealing the space between the cell membranes and cells [83]. As a result, long-chain PFASs are mainly limited to the root system and accumulate in the shoot system in a smaller amount [84,159,168].

PFAS concentrations in fruits decrease further due to additional barriers as the contaminants are translocated in the aboveground plant tissue [170]. It is important to note that uptake mechanisms are plant specific [83]. For example, Liu et al. linked differences in PFAS uptake in different crops to the varying lipid and protein contents [56]. Another study showed that the shoot accumulation of long-chain PFASs might be related to the surface-area-to-tissue ratio and is enhanced with an increasing surface area [171].

Uptake Rates of Different Crops

Since most plant uptake experiments have utilized varying conditions (e.g., soil vs. hydroponic cultivation, different plants or soil types, contamination levels, etc.), uptake metrics have been developed to improve the comparability of plant uptake across different studies. The bioaccumulation factor (BAF) gives the ratio of the PFAS concentration in the plant to the concentration levels in the surrounding environment (Equation (2)) [172]. The second important factor is the translocation factor (TF), which describes the distribution of the compound of interest within the plant (e.g., TFleaf/root) and is calculated similarly to the BAF (Equation (3)) [173].

BAF = [PFAS_{plant tissue}]/[PFAS_soil]

(2)

TF = [PFAS_leaf]/[PFAS_root]

(3)

The literature shows a direct positive correlation between the PFAS concentration in the soil and the plant [83,84,174]. Moreover, the bioaccumulation of PFASs by plants depends on various plant and soil factors. The physicochemical properties of the respective PFASs, contamination levels, plant type, and environmental conditions all interact to determine the uptake rate. An interesting trend observed was that, while TFs varied throughout the plant, BAFs were higher for vegetative plant parts than reproductive and storage organs [84,173]. The Casparian strip plays an important role in impeding the translocation within the plant, and translocation is also decreased by a relatively low protein content in the aboveground tissue [169].

Krippner et al. investigated the influence of the chain length (C4 to C10 for PFCA; C4, C6, and C8 for PFSAs) on the uptake and distribution of PFCAs and PFSAs in corn and observed that the uptake rates (BAF) exhibited a U-shaped trend. The lowest uptake rates were observed for PFHpA and PFHxS [175]. In contrast, the TF showed a negative correlation with the chain length for the tested PFCAs and PFSAs. Longer-chain-length PFASs possess a higher affinity for the lipid bilayer of membranes, which reduces overall movement within the plant [175]. PFOS showed the highest mean uptake rates, possibly due to different uptake mechanisms [176]. However, the BAFs for PFOS in different plants are generally considerably lower than those for PFOA or PFHxS [177]. It is apparent that translocation is more hindered with increasing chain lengths, and the risk of exposure is higher for short-chain PFASs that are less studied [178]. It should be noted that Krippner et al. grew plants in nutrient solutions, which is a less realistic approach since the interactions with soil components are missing [175].

In a case study, Liu et al. [56] examined the PFAS concentration in the soil and crops from a mega fluorochemical industrial park at two distances (0.3 km and 10 km). The soil had total contamination levels between 79.9 µg/kg and 200 µg/kg, with PFOA being the most abundant substance. The PFAS levels in the crops were between 58.8 µg/kg and 8050 µg/kg, with the BAFs of edible parts reaching levels of up to 48.0 µg/kg (radish) in shoot vegetables. The ranking for BAFs in terms of highest bioaccumulation of the different plants grown at the field 0.3 km from the pollution source were shoot vegetables (24.3 µg/kg) > fruit vegetables (6.63 µg/kg) > flower vegetables (4.23 µg/kg) > grain crops (4.05 µg/kg) > root vegetables (3.58 µg/kg).

PFASs’ impacts on plant health were examined by Stahl et al., who studied the uptake of PFOA and PFOS by different plants by spiking the soil with concentrations between 0 and 50 mg/kg. They found that exposure to a concentration of 50 mg/kg led to a significant reduction in corn (Zea mays) yield, mainly due to reduced ear mass [160]. However, these high contamination levels are only likely found at sites where PFASs are directly released (e.g., manufacturing or AFFF affected sites) [179], and studies investigating PFAS uptake at more typical environmentally relevant concentration levels are lacking [180].

3.4.2. PFAS Contamination in Livestock

The primary mechanism for PFAS contamination of livestock, and the resulting ani-mal-based productions they provide, is the consumption of contaminated feed or water. PFASs can be absorbed by the gastrointestinal tract [181,182] and bind to blood serum proteins [183,184]. The distribution in animal tissue is PFAS and species dependent. For example, PFOA, PFOS, and PFBS accumulate to a higher degree in the liver in most species, whereas PFBA and PFHxS can be found to a higher degree in the blood serum [182].

Kowalczyk et al. [25] investigated the kinetics of PFBS, PFHxS, PFOS, and PFOA in dairy cows by feeding them contaminated grass silage and hay for 21 days. The plasma concentration levels of PFBS (mean = 1.8 ±0.8 µg/L) and PFOA (mean = 8.5 ± 5.7 µg/L) were relatively low compared to those of PFOS (2462 ± 411 µg/L on day 44) and PFHxS (419 ± 172 µg/L on day 29). Additionally, the concentrations in the milk samples were proportional to blood serum levels, and the highest cumulative secretion was observed for PFOS (14 ± 3.6% µg/L). PFBS (0.01 ± 0.02% µg/L) and PFOA (0.1 ± 0.06% µg/L) were secreted into the milk in a low amount. Similarly, it was shown that PFOA is excreted significantly faster compared to PFOS in sheep. PFOA was mainly excreted with urine (51–55%) and could not be detected after day 42. In contrast, the highest PFOS excretion was detected in feces (4–5%) followed by milk. The PFOS levels in the tissue of the sheep did not decrease during the 21-day PFAS-free feeding period [26]. Since the animals were exposed to contaminated fodder for a relatively short time, serum and/or tissue concentrations might have not reached a steady state, possibly leading to a significant underestimation of these values [185].

In general, few studies have focused on PFASs in livestock, and their focus was on uptake and elimination. Potential adverse health effects were hardly investigated for different animal species, especially livestock [86]. Studies in monkeys showed a link between the exposure of PFASs and liver toxicity, altered thyroid hormone concentrations, decreased cholesterol serum concentrations, weight loss, and glycogen metabolism [186]. It is assumed that animals are capable of tolerating high PFAS concentrations [86], but more research is needed to investigate potential adverse health effects in livestock. Hlousova et al. found that PFAS concentrations from common farm livestock decrease in the following order: pig/bovine liver > egg > meat > dairy products (butter) [187]. Pasecnaja et al. summarized several studies that assessed the contamination levels in the European food market. They found that the most important sources were fish, meat, eggs, fruits, and vegetables, with fruits and vegetables exceeding the levels found in meat [188].

3.4.3. PFAS Transfer Through the Food Chain

Several parameters influence PFASs as they moves through the food production chain (Figure 4). The descriptions of the environmental fate of PFASs described above can be transferred to food production. Bioaccumulation can occur at every step if a contamination source is present. This is critical if the source continuously emits PFASs or if large amounts are released over a brief period. Crops grown on soil with high PFAS loadings will uptake the surrounding contaminants to an often times unknown extent. If these crops are used as feed, the livestock and the produced meat will be affected as well and might show stronger contamination levels compared to the fed crop. For example, Tefera et al. estimated the total daily mean dietary PFAS intakes for firefighters. They showed that food (i.e., eggs, fruits, and vegetables) which was produced at fire stations was responsible for 82% (i.e., 1250 ng/d) of the total PFAS intake [189].

The large number of individual PFAS molecules and their unique fates make it challenging to conduct highly accurate PFAS risk assessments [23] for food production systems. While short-chain PFASs have not been the focus of research, they show the highest mobility in the environment and across the food web, and their emissions are expected to increase. Recent studies showed that they might cause adverse health effects similar to long-chain PFASs [178,190,191], indicating the dire need for intensive investigations on shorter-chain PFASs.

There seems to be a disconnection between the fact that diet is a major exposure pathway [23] and the limited resources being used to address PFASs in agriculture. This might be partially due to early studies suggesting that dietary exposure is neglectable [192]. However, recent studies suggest that even exposure to an extremely low concentration might cause detrimental health effects, which is underlined by the EPA’s updated drinking water health guidance values [193] (see Figure 5).

The complexity of soil-to-plant, plant-to-human/livestock, and milk/meat-to-human PFAS interactions combined with the vast number of PFAS species limits our understanding of how and which PFASs are transferred through the food web. Another dimension of complexity is added when considering PFASs present in food contact materials (e.g., fast food wrappers or baking plates) [194]. This makes accurate assessments of the posed threat to consumers nearly impossible without more information.

3.5. Policy and Regulations

In the US, regulation and policy for pollution issues often occur at both the federal and state levels. There are currently no nationwide restrictions on PFAS contamination levels in food in the US. The State of Maine Department of Agriculture, Conservation and Forestry established PFOS action levels for beef (3.4 μg/kg) and milk (0.21 μg/L) [195]. The European Food Safety Authority set a tolerable weekly intake (TWI) for the sum of PFOA, PFNA, PFHxS, and PFOS at 4.4 (ng/kg food)/(week · kg bw) [196].

Most PFAS regulation and research funding have been directed at drinking water standards. Due to Minnesota’s longer history of PFAS pollution issues, Minnesota has had much earlier PFAS drinking water standards than the federal government. In the past few years, the federal government has begun to set more stringent requirements for drinking water (Figure 5).

Figure 5. A comparison of the proposed health advisory levels for PFOA (black) and PFOS (gray) drinking water standards by the Minnesota Department of Health (MDH) (continuous) and US Environmental Protection Agency (EPA) (dotted) between 2002 and 2024 [196,197]. Note that the scale is a log scale, so proposed standards for 2024 are on the order of 100,000 times more stringent than in 2002.

Figure 5. A comparison of the proposed health advisory levels for PFOA (black) and PFOS (gray) drinking water standards by the Minnesota Department of Health (MDH) (continuous) and US Environmental Protection Agency (EPA) (dotted) between 2002 and 2024 [196,197]. Note that the scale is a log scale, so proposed standards for 2024 are on the order of 100,000 times more stringent than in 2002.

Until recently, neither the state nor federal government had addressed PFAS contamination issues in agriculture. In 2023, the Minnesota Legislature passed bills (SF 1955 and HF 2310) regulating pesticide products that contain intentionally added PFASs [197]. One of the new laws requires pesticide registrants to inform the Minnesota Department of Agriculture if a pesticide product contains intentionally added PFASs. The EPA is studying this issue and determining the scope and scale of pesticide-related PFAS issues.

Both Minnesota and the federal government have set up near-term plans for further research and rulemaking on PFAS-related issues. Minnesota’s PFAS staff group has observed that it is challenging to manage PFASs sufficiently through regulatory actions as several vital areas (e.g., pollution prevention and waste management or understanding risks and how they relate to exposure through food/water) are overlapping [198]. The EPA’s roadmap’s goals for 2021 to 2024 are to increase the understanding of the effects of PFAS exposure on human health and ecological systems, prevent the release of those sub-stances into the environment, and facilitate remediation efforts at contaminated sites [199].

In Minnesota, the MPCA, MDH, Department of Natural Resources, and Minnesota Department of Agriculture published Minnesota’s PFAS Blueprint for addressing the PFAS problem systematically and efficiently at the beginning of 2021.

Minnesota’s blueprint lists the following areas of concern: (1) preventing PFAS pollution, (2) measuring PFAS effectively and consistently, (3) quantifying PFAS risks to human health, (4) limiting PFAS exposure from drinking water, (5) reducing PFAS exposure from consuming fish and game, (6) limiting PFAS exposure from food, (7) understanding risks from PFAS in the air, (8) protecting ecosystem health, and (9) remediating PFAS-contaminated sites [198].

While all agree that the potential for PFAS contamination in food production needs to be addressed [83], regulating agriculture is often made more difficult due to political and practical issues. Minnesota’s Department of Employment and Economic Development estimated that the agricultural sector in Minnesota generated about USD 17 billion in sales, with USD 8.85 billion being contributed from the cultivation of crops in the year 2022 [200]. Similar economic impacts are generated by agriculture in states throughout the US Midwest. Agriculture is an important industry that supplies the planet with needed food; thus, state and federal regulators are often cautious about changing rules for agriculture. However, care needs to be taken to reduce the potential that agricultural products are a source of exposure to PFAS in people. Poorly addressing PFAS issues has the potential to negatively impact human health, threaten economic sustainability, and cause consumer backlash.

The PFAS cycle must be broken long term to prevent further contamination on agricultural sites. This requires far-reaching regulations governing the production, use, and disposal of PFASs as it is very challenging to prevent the emission of PFASs into the environment over their life cycle [201]. Therefore, restricting production and utilization is the best option to prevent further contamination. For example, the state of Minnesota banned the use of PFAS-containing firefighting foam with a few exceptions and limited intentionally adding PFASs to food packaging material in 2024 [202]. Banning the commercialization of PFAS-containing products is especially important if the PFAS-provided properties are not necessarily required for the respective application. This step must also include thorough and mandatory testing of PFAS substitutes to ensure their biocompatibility and biodegradability. Biosolids for land application should be thoroughly investigated, and PFAS contaminants exceeding a set threshold should be treated as they are primary contamination sources in rural areas. The Maine legislature went so far as to ban land application in August 2022 [203]. Existing wastes should be handled adequately so they will not enter the groundwater or atmosphere. Furthermore, law makers should address PFAS-containing food contact material and restrict their use if necessary. Consequentially, legally binding maximum PFAS levels in food should be established, which should include the most common legacy PFASs (e.g., PFBA, PFBS, PFHxA, PFxS, PFOA, PFOS, and fluorotelomer sulfonates) and upcoming alternatives (e.g., GenX, F-53B, OBS).

Lastly, agencies should facilitate mitigation strategies, which are still energy- and cost-intensive. It is self-explanatory that the required actions require enormous resources and the parties responsible for the pollution to be held accountable.

3.6. Future Research Recommendations and Challenges

It is clear that the US agriculture community faces challenges in identifying the risks that PFASs present to today’s agriculture-based food systems. More work across all areas is needed to understand the scope of these challenges and develop long-term strategies for mitigating related risks. It is also important to recognize that there is a difficult question of responsibility for PFAS contamination as farms are most likely contaminated without the knowledge of the farmers. However, they may suffer financial losses due to contamination or mitigation. A rapid response and sufficient resources are needed for establishing a framework to tackle the PFAS crisis.

3.6.1. Agricultural Challenges and Strategies for PFAS Mitigation

Arguably, most US farmers have not heard of the term PFAS. Thus, a knowledge gap in the farming community is an initial barrier to mitigating risks to agriculture and the food supply. Little guidance and studies in the literature are available that are geared towards informing agricultural practitioners about PFASs. The agriculture sector is already under heavy pressure to act on other environmental issues (e.g., greenhouse gases, nitrates, fossil fuels, and pesticides), which makes it difficult to incorporate additional information about PFASs in outreach materials to producers and their cropping advisors. Without this information, farmers are unlikely to even be aware that they should consider potential risks to their crop or livestock products from PFASs.

It is also clear that the outlined knowledge gaps need to be addressed regarding translocation in agricultural plants and livestock. Therefore, more research on PFAS translocation and presence is needed in every step of food production (soil sorption, plant uptake/translocation, transfer to livestock, and consumption). In many products, long-chain PFASs have been replaced with shorter alternatives (e.g., GenX), which are more mobile in agricultural systems. However, there are uncertainties about their toxicological potential that must be studied. Field studies that deal with plant uptake are needed to enhance the understanding of the respective uptake mechanisms. Clinical studies that expose the livestock to relatively high PFAS loads have limited applicability as well, which makes it challenging for regulators to establish suitable action plans. One helpful measure would be to develop TWI values to increase the transparency and trust of consumers.

These suggested efforts require immediate action, many resources, and the bundling of expertise across biological, chemical, and medical fields as the question of under which conditions crops and livestock can be safely grown must be answered.

Finding data on PFAS-contaminated agricultural sites that present risks to farm operations is also difficult. States with robust environmental protection policies or a known existing issue with PFASs may have a well-developed database on PFAS manufacture, use, or contamination sites. For example, Minnesota has a repository of PFAS-related test data available online in raw and searchable (map based) formats [50]. A further challenge for producers that have identified a potential PFAS contamination risk is the lack of resources on subsequent steps to verify the risks. Educational and consulting resources for farmers in the US are typically centered on state universities and their federally funded agricultural extension staff. Extension staff in some states have begun outreach efforts on agricultural PFAS issues [204], typically in states with known significant PFAS problems. However, the majority of state universities do not have staff focused on the issue. Farmers at high risk for PFAS contamination will likely need to refer to information from other states or outside the agricultural sector.

Should a producer believe that they may be at risk for PFAS issues in their operations, the next step would be testing. The availability of testing labs and methods needed for performing soil/plant/animal testing will be another challenge for farmers. Tests for PFASs are very sensitive, detecting these materials at the parts per billion or trillion level. A simple web-based search indicates that most certified labs testing drinking water samples in the US charge between USD 250 and USD 600 per sample. It is likely that testing soil or products would add additional costs. Between the costs of tests and sampling work, it is likely that farmers would find testing unaffordable. Moreover, financial support should be provided for contaminated farms so that farmers do not fear testing. It may be necessary for states to assist with the expertise and cost of conducting PFAS tests. There is also the issue of interpreting test results as there are currently no guidelines for defining soil contamination in an agricultural context.

3.6.2. Strategies for PFAS Risk Reduction on Farms

Understanding the sources and associated entry of PFASs into the environment is crucial to identify locations with higher associated risks and to develop mitigation strategies to ensure efficient use of resources [205]. Based on the findings in Minnesota, these include rural sites located by landfills or airports or sites that have undergone application of contaminated biosolids. In the case of landfills and airports, these are typically matters of public record. Farmers who have been applying biosolids will typically know how much and how often they have applied them to their land. This work can all be performed at the individual farm level or by mapping farms adjacent to known contamination sites.

Another approach to identifying risks of PFASs in agriculture that may play an increasing role in the future is machine learning (e.g., for farms in proximity to airports or military sites). Maine is already examining machine learning-related methods [195]. However, extensive PFAS and environmental monitoring data is required for these methods to be efficiently applied.

Once identified, sites with elevated contamination levels or at high risk for new contamination must be further evaluated to assess whether and how farming can be continued. Remediating agricultural sites is currently not economically feasible due to the energy and cost-intensive nature of the respective methods. Alternative crops that have lower BAFs or are not used for food production (e.g., bioenergy or fiber crops) may need to be produced on contaminated land. PFASs would still be present on both land and crops but would not put livestock or humans at risk.

Another contamination-related issue is irrigation and groundwater. Agricultural lands may not necessarily be heavily contaminated, but irrigation water pumped from contaminated aquifers could pose a risk. Again, the primary rural source of contamination for water is landfills; thus, it is prudent to test irrigation water being pumped from wells near landfills. This should also be performed for drinking water provided to livestock.

At present, these are concrete steps that can be taken to mitigate PFAS risks. They will likely need to be re-evaluated and updated as more information becomes available on both PFASs in general and its linkage to agriculture in particular. The rapid development of strict drinking water standards suggests that PFAS regulation will be developed for agriculture in the near future so that we can reduce the potential for PFAS contamination in food supply.

4. Conclusions

The experience with PFASs in Minnesota makes it evident that PFAS contamination in agriculture is an area of concern that needs to be more fully examined by the scientific community, policymakers, and farmers to mitigate risks for the environment and consumers. Minnesota, with its historical relationship to PFASs and a strong agricultural sector, provides a glimpse into the major sources of agricultural contamination (landfills, biosolids, and airports) and can be used as a model region for other rural areas. Adaptations must be made based on local geographical features, its contamination history, and individual food production chains.

Knowledge gaps are a critical issue, as consuming PFAS-contaminated foodstuffs is identified as an important exposure pathway. Our findings suggest that more field studies which investigate the PFAS partitioning behavior in all soil types, PFAS uptake by crops cultivated in different soil types, and PFAS uptake/distribution/accumulation in livestock are direly needed. Studies in a controlled environment can expand the understanding of certain factors but can only be applied to real conditions to a limited extent. Often times, environmental conditions are not investigated, but unrealistic high PFAS levels are used. Epidemiological studies are lacking in this area, as well as data and reports specifically evaluating contamination levels at agricultural sites. The amount and type of specific PFASs detected in the environment will change in the near future due to manufacturing adaptions made by the industry. This poses a great challenge to the scientific community and must be tackled quickly so that research, regulations, and public health measures do not only address legacy PFASs.

The high complexity, knowledge gaps in critical areas, and interdisciplinary nature of the PFAS crisis do not allow reliant risk assessments at the present time. No simple approaches will adequately meet this problem, and complex approaches, such as machine learning or intergovernmental cooperation, will be of utmost importance.

Our findings suggest that the consumption of produce from areas with no direct PFAS application should pose no significant threat to the consumer. Animal products, especially offal, generally show higher contamination levels compared to plant-based foods. It is advisable to establish a monitoring system in agricultural areas and across the food chain. Farms and foods with a higher associated risk of contamination (e.g., close to manufacturing sites, airports, or landfills; those that have undergone historic biosolid application; or those in contact with contaminated surface water) should be preferred during the planning phase of such systems. Legally binding regulations dealing with PFASs in agriculture and food production are basically non-existent, which shows the need for immediate action.

The current strategies to limit agricultural PFAS contamination rely on avoiding known contamination from these sources or leachates spreading in groundwater. Remediation efforts will likely play a significant role in the future for some agricultural sites as the methods become economically competitive or are required by future regulations. However, there are currently no sensible remediation strategies in the context of agriculture. Established methods, such as the excavation or washing of contaminated soils, do not meet the requirements, and more attention needs to be paid to this aspect.