PFAS in Agricultural Systems: Guidance for Conservation Programs at USDA (2026)
Chapter: 5 Applied Research Gaps for PFAS Management in the Context of Conservation
5
Applied Research Gaps for PFAS Management in the Context of Conservation
To solve problems related to per- and polyfluoroalkyl substances (PFAS), research is underway on many fronts, including on treatment technologies that safely remove and destroy PFAS and for replacement substances that have equivalent functionality but do not harm human health or the environment. In the context of conservation on the land, applied research needs to focus on minimizing PFAS uptake into plants and animals, in situ sequestration, and removal of PFAS to the greatest extent possible. This chapter reviews four areas of research that could advance the ability of conservation practices to address PFAS contamination on agricultural land: a better understanding of PFAS fate and transport in different types of soils, mechanisms by which to trap or sequester PFAS, an improved understanding of PFAS uptake in plants, and an improved understanding of PFAS uptake in animals.
DISCERNING PFAS FATE AND TRANSPORT IN VARYING SOIL TYPES ACROSS THE UNITED STATES
As described in Chapter 2, no single factor can determine PFAS fate and transport within and through soils (Li et al. 2018; Wang et al. 2023). A multitude of soil factors (e.g., clays/oxides, the amount and type of soil organic matter, pH, the presence of various cations, soil texture, soil–water relationships) affect PFAS fate, transport, and risk to producers, farms, and surrounding ecosystems. Climatic conditions (e.g., precipitation, wind, temperature) also must be considered (Box 5-1).
Furthermore, the fate and transport of the various PFAS types present in soils need further evaluation (e.g., zwitterionic, cationic, or anionic PFAS; short- versus long-chain PFAS). For example, Wang et al. (2023) pointed out that zwitterionic PFAS research is lacking and that soil pH likely changes the zwitterionic PFAS charge, thus affecting sorption. The authors also stated that PFAS sorption changes between mineral
BOX 5-1
Fate, Transport, and Climate
One of the major differences between greenhouse and field studies of PFAS in soils is the effect of climate. For example, higher temperatures increase plant metabolism and transpiration and thus increase PFAS transport and deposition in plant parts (i.e., leaves). An increase in temperature (from 20 °C to 30 °C) has been shown to increase PFAS concentrations by about 2-fold in wheat, with short-chain PFAS accumulation less sensitive to temperature changes (Zhao et al. 2016). Although the study by Zhao et al. (2016) was performed in the laboratory, it sheds light on the effect of field temperature on plant PFAS uptake.
Additionally, precipitation plays a major role in PFAS fate and transport. To the committee’s knowledge, there have been no long-term studies linking PFAS fate and transport to variable precipitation patterns—thus, a knowledge gap exists. Often in pot studies with plants, leaching is kept at a minimum and thus short-chain PFAS have a greater likelihood of contacting roots, unlike what may occur under field settings (Costello and Lee 2024). Lasters et al. (2024) attempted to link PFAS bioavailability in gardens to soil factors but found it nearly impossible to link the two together. The authors suggested that perhaps soil porewater (which was not quantified) may better indicate greater PFAS availability in gardens and thus potentially in field settings. Unlike greenhouse trials, field studies need to identify site-specific PFAS sources, focus on a larger number of PFAS and precursors, keep a close eye on site climatic conditions that affect soil conditions (perhaps via the use of weather stations, dataloggers, and soil moisture and temperature sensors), and assess PFAS risks to water sources (Costello and Lee 2024).
types, and whether sorption dominates on clay surfaces versus within clay interlayers depends on clay type (e.g., kaolinite versus montmorillonite; Zhang et al. 2014). A better understanding of the relationships between various clays and PFAS would help to advance knowledge of PFAS sorption mechanisms affecting fate and transport in soils (Mejia-Avendaño et al. 2020). Moreover, it is understood that short-chain PFAS do not readily sorb to soil phases compared with long-chain PFAS. Therefore, Nasrollahpour et al. (2025) pointed out that further sorption or removal research (e.g., clays, other soil phases, or other materials used for remedial purposes) should focus on short-chain PFAS as they are increasingly used to replace long-chain PFAS. Wang et al. (2023) provided a concise starting point for understanding the factors that affect PFAS sorption in soils exemplified for perfluoroalkyl sulfonic acids (PFSAs) and perfluoroalkyl carboxylic acids (PFCAs; Table 5-1). Additional factors to consider in future research are soil texture (Li et al. 2018; Nguyen et al. 2020; Mei et al. 2021; Umeh et al. 2021) and iron and aluminum oxides (Umeh et al. 2021; Campos-Pereira et al. 2023).
TABLE 5-1 Factors that Impact Sorption of PFCAs and PFSAs on Soil
| Factor | Effect on Sorption | Research Finding | ||
|---|---|---|---|---|
| Category | Specification | on the Effect | Reference | |
| PFAS Property | Functional Groups | PFSA > PFCA | Consistent | Gellrich et al. 2012; Zhao et al. 2014; Milinovic et al. 2015; Loganathan and Wilson 2022; Luft et al. 2022; Wang et al. 2022 |
| Carbon Chain Length | Longer > Shorter Chain Length | Consistent | ||
| Soil Property | Organic Carbon (OC) | Higher > Lower OC | Mixed | Becker et al. 2008; Kwadijk et al. 2010; Pan and You 2010; You et al. 2010; Ahrens et al. 2011; Chen et al. 2012; Milinovic et al. 2015; Luft et al. 2022 |
| Cation Exchange Capacity (CEC) | Higher > Lower CEC | Mixed | Barzen-Hanson et al. 2017; Xiao et al. 2019; Mejia-Avendaño et al. 2020; Nguyen et al. 2020 | |
| Anion Exchange Capacity (AEC) | Higher > Lower AEC | Mixed | Barzen-Hanson et al. 2017; Li et al. 2019 | |
| Humic Acid (HA) | Lower > Higher HA in Low Soil Content | NDa | Zhang et al. 2014 | |
| Higher > Lower HA in High Soil Content | Higgins and Luthy 2006 | |||
| Minerals | Positively Charged > Negatively Charged Minerals | NDa | Hellsing et al. 2016 | |
| Montmorillonite > Kaolinite | Zhang et al. 2014 | |||
| Factor | Effect on Sorption | Research Finding | ||
|---|---|---|---|---|
| Category | Specification | on the Effect | Reference | |
| Water Quality Property | pH | Lower > Higher pH | Mixed | Higgins and Luthy 2006; Kwadijk et al. 2010; Kwadijk et al. 2013; Martz et al. 2019; Oliver et al. 2019 |
| Cation | Higher > Lower Divalent Cation Concentration | Consistent | Schwarzenbach et al. 2002; Higgins and Luthy 2006; Kwadijk et al. 2010; You et al. 2010; Chen et al. 2012; Cai et al. 2022 | |
| Higher > Lower Monovalent Cation Concentration | Mixed | Higgins and Luthy 2006; Chen et al. 2013; Cai et al. 2022 | ||
a ND = No basis to decide because only one study reports.
NOTE: AEC = anion exchange capacity; CEC = cation exchange capacity; HA = humic acid; OC = organic carbon.
SOURCE: Wang et al. 2023.
Although there has been considerable work conducted on PFAS sorption, less is understood about PFAS desorption hysteresis, which greatly impacts the behavior of legacy PFAS (perfluorooctanoic acid [PFOA] and perfluorooctane sulfonic acid [PFOS]). The limited desorption data that are available are inconsistent, which is often driven by desorption results from high-concentration lab-spiked soils that showed little desorption hysteresis (Umeh et al. 2024) versus historically contaminated field soils (Schaefer et al. 2022; Klamerus et al. 2025). For the latter, 50 to almost 80 percent of PFOS in soil resisted desorption after several sequential desorption steps (Schaefer et al. 2022; Klamerus et al. 2025). A benefit to sorption, desorption, and hysteretic studies would be to combine them with modeling approaches to further understand PFAS behavior in the environment. The majority of PFAS modeling has focused on groundwater, while less attention has applied modeling approaches to predict PFAS concentrations in soil, as outlined in Chapter 4. More research is needed in this space.
Better insight is needed into how soil and climatic factors coupled to PFAS characteristics affect PFAS fate and transport across the United States for a more in-depth understanding of PFAS risks across the country. To address these research gaps, a national network of researchers could be created to systematically study PFAS fate and transport with a focus on soils containing varying constituents (e.g., clay types or oxides) within individual U.S. regions (e.g., based on soil survey regions, major land resource areas, and climate). Researchers could use the predictive modeling tools
reviewed in Chapter 4 to locate potentially affected soils, identify soil characteristics based on the Natural Resources Conservation Service’s (NRCS’s) Web Soil Survey in context for factors that affect PFAS sorption on soils (Table 5-1), and consider already conducted field studies from which data gaps and future needs can be identified. Results could be used to expand Table 5-1 and could be added to the data fields in the Web Soil Survey for PFAS attenuation in soils and PFAS movement classes (see Box 4-3 in Chapter 4). Research findings then could be directed toward addressing the positive and negative effects of conservation practices when utilizing this newly found information (e.g., expanding upon Table 3-2 in Chapter 3). This same type of national research group approach could be utilized to target other research gaps noted below.
OPPORTUNITIES TO TRAP OR SEQUESTER PFAS
Means by which PFAS can be sorbed and sequestered in soils to reduce their bioavailability are now being studied. The benefits to sequestering PFAS include the potential to reduce plant uptake as well as leaching. Potential disadvantages are that PFAS are held in place, which may hinder success of future removal or destruction strategies. There is also the potential for changes in the PFAS sequestering sorbents over time that may lead to unforeseen release of PFAS. If sorbents are to be used effectively, more research is desperately required given the multitude of compounds in the PFAS family. Additionally, if PFAS are an issue in U.S. waters, finding the means by which PFAS can be removed from such water bodies is of paramount importance. The section below outlines possible approaches in targeting PFAS sorption within soils or waters.
Potential Sorbents for PFAS Sequestration on Agricultural Land
Various sorbents, such as biochars, modified clays, activated carbons, various drinking water treatment residuals (DWTR), nanoparticles, and even wood chips blended with biosolids have been studied for their PFAS-sorption affinities. An overview of recent research focused on this topic is outlined below to provide thoughts for filling research gaps.
Biochars
Several designer sorbents have been tested at the bench scale to target high PFAS sorption capacity; however, scalability in terms of provision and cost at the landscape scale are unlikely. Therefore, much attention has turned toward biochar that can be produced in large volumes. Biochars are created by pyrolyzing organically based feedstocks in an oxygen-depleted or oxygen-devoid environment. Pyrolysis temperatures often range from 350 °C to 700 °C, although greater temperatures have been utilized for biochar creation. Biochars have been proven to be useful sorbents for heavy metals and trace organic compounds from waters and soils, and thus a great deal of interest is presently focused on biochar use as a sorbent for PFAS. Biochar is specifically named as a candidate amendment under the Soil Carbon Amendment conservation practice
standard (Code 336; USDA 2022) to improve soil health. However, at present, its ability (as well as the ability of other possible sorbents) to sequester PFAS is not part of a practice standard to address contamination.
A recent review of using biochars for mitigating PFAS in agricultural settings highlighted important biochar characteristics for PFAS sorption (Ramos and Ashworth 2024). First, biochar can be effective at sorbing PFAS, but success with respect to short-versus long-chain PFAS depends on feedstock and pyrolysis conditions. For example, Inyang and Dickenson (2017) showed that both softwood (i.e., spruce pyrolyzed at 650 °C) and hardwood (source unknown, gasified at 900 °C) sorbed long-chain PFOA equal to activated carbon, yet only hardwood approached short-chain perfluorobutanoic acid (PFBA) removal as compared to activated carbon (87 percent versus 94 percent sorption, respectively); softwood biochar only sorbed 18 percent of PFBA. Inyang and Dickenson (2017) noted that the PFBA sorption efficiency of biochar appeared to increase within increasing surface area. Liu et al. (2021) compared pyrolysis temperature (500 °C, 700 °C, and 900 °C), hold time (3, 5, and 6 hours), and different feedstocks (corn cob, common reed [Phragmites australis], aspen wood chips, and soybean dreg) on biochar PFOA and PFBA removal efficiency; results are presented in Figure 5-1. PFOA removal efficiency for corn cob, reed straw, and aspen chip biochars were around 100 percent, similar to activated carbon, while soybean dreg was inferior. Removal efficiencies for PFBA by corn cob and reed straw biochar were 65 percent and 85 percent as compared to activated carbon (around 15 percent); aspen chip and soybean dreg for PFBA removal were lower than activated carbon. Pyrolysis temperature impacted PFOA and PFBA removal efficiency; increasing temperature to above 700 °C led to 100-percent removal of PFOA, yet PFBA removal efficiency was negligible at 700 °C. Not until pyrolysis temperature reached 900 °C was PFBA removal efficiency elevated to around 80 percent.
Most literature on pyrolysis, including that cited here, determine removal efficiency based on removal of the parent compound targeted. It is possible that shorter-chain PFAS were generated, including ultra short-chain trifluoroacetate (TFA) and perfluoropropanoic acid (PFPrA); a more complete picture is determined using a fluorine mole approach. Pyrolysis hold time also played a role, and as hold times increased from 3 to 5 to 6 hours, so did PFOA removal efficiency (80 to 95 to 96 percent, respectively). It is important to note that hold times are often not described in the biochar literature; however, the optimal hold time for creating biochar to remove PFBA was 5 hours, as efficiency decreased with a hold time of 6 hours. The explanation as to why efficiency decreased could have been due to a reduction in pore size diameter with increasing hold time; in general, pores need to be approximately two to three times the molecular diameter of the PFAS in order to be trapped (Zimmerman et al. 2004).
Second, feedstock and pyrolysis conditions that alter specific surface area and biochar carbon–oxygen ratios play a role in PFAS sorption. A subset of data obtained by Ippolito et al. (2020) targeting biochar properties from feedstocks potentially used for agricultural purposes (wood-based, crop waste, manure/biosolids) was used by Ramos and Ashworth (2024). Ramos and Ashworth (2024) then scoured the literature for information on biochars made from those feedstocks and their ability to sorb
NOTE: Removal efficiency was discerned based on the disappearance of the target perfluoroalkyl acids (PFAAs) after adsorption over 24 hours with an initial concentration of 100 µg/L and an adsorbent dose of 0.2 g/L. GAC = granular activated carbon.
SOURCE: Liu et al. 2021.
PFAS. A principal components analysis was performed to identify linkages between PFAS removal and biochar feedstock type, pyrolysis temperature, and a variety of characteristics. The authors found that as a function of feedstock type, the greater the specific surface area (SSA) and the greater the carbon–oxygen ratio, the greater PFAS removal. Increases in SSA are most often related to increases in pyrolysis temperature. Increases in the carbon–oxygen ratio are often associated with greater hexane-like carbon structures formed during pyrolysis (via increased temperature) leading to an increased presence of Π-electrons that can sorb PFAS, along with an increase in hydrophobic carbonaceous adsorbents (Gagliano et al. 2020). Ramos and Ashworth (2024) did not specifically describe feedstock influence on PFAS sorption, but according to Ippolito et al. (2020), SSA increases from wood to crop to manures/biosolids-based feedstocks. Also noteworthy is that, with a greater carbon–oxygen ratio, biochar has a greater likelihood of having a half-life of more than 1,000 years (Ippolito et al. 2020), increasing its ability to potentially sequester PFAS long-term. Research over time (i.e., over the course of many years) is needed to verify this contention.
Modifications in biochars have been performed to potentially enhance their ability to sorb PFAS. Wu et al. (2022) modified switchgrass, water oak leaves, and biosolids feedstocks with either ferric chloride (FeCl3) or carbon nanotubes prior to pyrolysis and biochar creation. Carbon nanotube modification had little effect on PFAS sorption, yet the authors found that PFOA sorption onto all biochars was approximately doubled when modified with FeCl3 (as compared to no modification). PFOA sorption onto FeCl3-modified biochars followed this order: water oak biochar (102 µmol/g) < switchgrass biochar (112 µmol/g) < biosolids biochar (470 µmol/g). Wu et al. (2022) emphasized that understanding biochar metal content, pore volume, surface area, and surface functional groups present all play important roles in sorbing PFOA via a combination of electrostatic, physical, and hydrophobic interactions. Rodrigo et al. (2022) added magnetite (Fe3O4) to Douglas fir biochar, noting that PFOA sorption increased from 9 (unmodified biochar) to over 650 mg/g. Sørmo et al. (2021) created wood-based biochars either without or with increasing molar ratios of steam or carbon dioxide (CO2) to feedstock carbon. Increasing amounts of biochar were then added to PFAS-contaminated soils containing either relatively low or high soil organic carbon content. At a 5-percent amendment application rate, all biochars almost completely reduced PFAS leachate concentrations (98–100 percent) in the low organic carbon-containing soil; in the high organic carbon-containing soils, PFAS leachate losses were variable (23–100 percent). Increases in PFAS sorption onto biochar was attributed to increasing biochar internal surface areas and porosity associated with treatment modifications. Sørmo et al. (2021) pointed out that although biochar modifications were successful in increasing PFAS sorption, the amount of biochar end product was reduced after modification, which is something to consider when performing steam or CO2 modification. Others have used varying modifications (e.g., alkali solutions) to increase PFAS sorption (e.g., Zhou et al. 2021). Overall, it is important to note that any post-biochar-creation modification will add cost to the final product and may cause loss of biochar during the process, which must be considered in choosing biochar modifications for PFAS sorption.
Last, a great deal of research has focused on research-grade (lab-modified) biochars for PFAS sorption and remediation. Future work should devote greater attention
on PFAS sorption from production-scale biochars, potentially suggesting alterations in pyrolysis conditions (e.g., temperature, hold times) if the goal is to increase PFAS-sorption efficiency (i.e., high capacity, fast sorption rates, and limited desorption potential). Perhaps a suite of biochars, based on temperature and hold time, could be created from one feedstock in order to target both short-chain and long-chain PFAS sorption. Attention could also focus on the creation of biochars on-site or near-site from PFAS-contaminated plant materials, keeping in mind the feedstock and pyrolysis conditions, barring concerns on incomplete PFAS combustion in the process and air emissions, to be returned to on-site contaminated soils for PFAS sequestration and reduction in bioavailability. On- or near-site biochar production could potentially reduce overall cost of biochar.
Future work should also focus attention on the use of biochar in a greater number of field settings, as most research has been performed within laboratory-controlled environments. Results from ongoing field studies show variable responses across multiple harvests and years; these results could be due to factors such as sorption–desorption dynamics, mode of application, depth of incorporation, moisture regime, competition from other chemicals in soils for sorptive sites by biochars, and crop species. Depth at which the initial carbon amendment is incorporated will depend on the depth of discing or tilling. However, particles can move during the growing season due to root growth and water movement. Soil type also affects movement; for example, sandy soils and soil profiles with significant preferential flow paths have a higher probability of much greater vertical transport of PFAS-laden particles. Movement deeper into the soil profile can lead, over time, to particle movement to groundwater or to tile drains that are discharged to surface water bodies. As discussed above (see section “Discerning PFAS Fate and Transport in Varying Soil Types Across the United States”), strong desorption of hysteresis of PFAS has been observed for historically contaminated field soils, particularly those contaminated with long-chain PFAS. No such work has been conducted on biochar-amended soils. Therefore, field studies and some supporting bench-scale studies are needed to quantify sorption–desorption hysteresis for varying PFAS alkyl chain length from biochar-amended field soils, assess in-field long-term (e.g., years) effectiveness, evaluate the potential benefits of repeated applications, and account for soil types and properties and climatic conditions.
Quantifying these characteristics applies to any sorbent (including the clays and DWTRs discussed below). However, given the particularly high level of interest in biochars as PFAS sorbents, taking advantage of machine learning and artificial intelligence is warranted to guide the decision-making process in biochar choices and optimization for mitigating PFAS (Nasrollahpour et al. 2025).
Modified Clays
As briefly discussed in Chapter 2, negatively charged clays can attract divalent cations that lead to cation bridging between clays and negatively charged PFAS or directly bind positively charged PFAS (e.g., Munoz et al. 2018; Mejia-Avendaño et al. 2020). Smectitic clays, which have greater surface area for reactivity, are typically
more attractive for sorbing PFAS compared with kaolinitic clays (although PFAS charge plays an important role). However, this is not always the case, as others have noted kaolinite sorption of PFAS to be greater than smectitic clays (e.g., Zhang et al. 2014; Zhao et al. 2014). Attention has focused on the use of high surface area clays (either by themselves or via some modification) or low surface area clays to alter PFAS fate, transport, and bioavailability.
Hearon et al. (2022) utilized montmorillonite clay (i.e., high surface area clay) by itself or amended with either carnitine or choline (added to increase hydrophobicity). Clay or modified clay was applied at 2 percent to a PFAS-spiked soil (1 µg/g) and a vegetative growth toxicity assay was performed using common duckweed (Lemna minor). The two modified clays reduced PFOA and PFOS bioavailability by about 58 percent and 78 percent, respectively, as compared to a control. The unmodified montmorillonite clay was not as effective as the modified clays, yet it still reduced PFOA and PFOS bioavailability by about 45 percent and 70 percent, respectively. Hearon et al. (2022) also focused attention on clay or modified clay (at 2 percent) application to a 1:1 mixture of soil:compost, introducing the soil to a solution containing nutrients and 1 µg/mL PFAS and using the soil to grow cucumber (Cucumis sativus). Across all plant parts (e.g., roots, stems, and leaves), the clay, carnitine-modified clay, or choline-modified clay reduced PFOA translocation by 51, 63, and 64 percent, respectively, while reducing PFOS translocation by 50, 70, and 67 percent, respectively. Hearon et al. (2022) suggested that modified clays might not only help with remediation strategies but help reduce PFAS uptake into certain plant components.
Others have focused on the use of kaolinitic clay (i.e., low surface area) to sorb PFAS. For example, using 1 g/L PFOS and 5 g/L clay, Zhang et al. (2014) showed that kaolinite outperformed montmorillonite for sorbing PFOS with 78 µg/g versus 54 µg/g sorbed, respectively. Higher sorption of PFAS to kaolinite was attributed to kaolinite being an electrically neutral mineral with part of it having a hydroxyl surface that may lead to greater PFAS sorptivity. The latter conclusion was also noted later by Ke et al. (2023) in their PFOS sorption studies.
The above findings, however, need to be taken with some caution. First, the ability of inherent soil clays to sorb anionic PFAS may be weak unless cation bridging occurs, or they may be completely ineffective at sorbing hydrophobic PFAS as these clays are hydrophilic due to hydration of cations sorbed on exchange sites of clay particles (Bolan et al. 2021). Thus, modifying clays (external to soil) with a material or surfactant can increase its hydrophobicity, leading to increased hydrophobic PFAS sorption (Guégan 2019) post soil application.
Second, over time, (modified) clays added to soils likely will react with organic matter and competing ions and will be affected by soil pH, ionic strength, and temperature (e.g., Mukhopadhyay et al. 2021), potentially reducing the effectiveness of the modified materials to sorb PFAS (Figure 5-2). Thus, future studies should focus on long-term investigations using (modified) clay mineral mixtures for enhanced sorption and reduction in PFAS fate, transport, and bioavailability, keeping a close eye on on-site climatic and soil conditions. To help guide this area of study, one could start with results presented in Table 5-2 from Mukhopadhyay et al. (2021).
SOURCE: Mukhopadhyay et al. 2021.
Drinking Water Treatment Residuals
Drinking water treatment processes generate a variety of waste products depending on source water, chemicals used for clarification/purification, and other operations utilized (Ippolito et al. 2011). The spent chemicals created post water clarification, DWTRs, typically contain aluminum, iron, or calcium from chemical treatment, in addition to any minerals (e.g., clays, oxides) removed from the source water. Because of the chemical makeup of DWTRs, and the amount generated in the United States on a daily basis (around 2 million tons; Broadbent et al. 2025), they may potentially be a low-cost PFAS sorbent. However, it is important to note that the chemical make-up of DWTRs is heterogenous, a function of suspended sediment in the water column entering the water treatment facility. Hence, the use of DWTRs for sorption of PFAS at a particular location may or may not be optimized via the use of local DWTRs. Resident PFAS on DWTRs, which will vary with water sources and treatment processes, must also be considered. Gravesen et al. (2025) found perfluoroalkyl acid concentrations ranged from 0 µg/kg to 3.3 µg/kg in seven tested DWTRs. Therefore, even though they are inexpensive, more research is needed to determine that DWTRs used as sorbents are optimized to the site of application and are not a source themselves of PFAS.
Zhang et al. (2021) were among the first to report on the use of DWTRs for PFAS sorption. An aluminum-based DWTR, the result of using alum salts in water treatment, was studied for its PFOA and PFOS sorption capabilities from solution. Across a range of solution pH values (3 to 11), aluminum-based DWTRs practically sorbed all PFAS
TABLE 5-2 Physical and Chemical Properties of Clay-based Adsorbents for PFAS Removal
| Clays and clay minerals | Physical properties | Chemical properties | Reference |
|---|---|---|---|
| Natural clays and clay minerals | Surface area: montmorillonite = 67.5 m2/g, kaolinite = 23.1 m2/g, hematite = 9.9 m2/g; Porous structure. | CEC: montmorillonite = 111 cmol/kg, kaolinite = 34 cmol/kg, hematite = 78 cmol/kg; Net negative surface charge due to isomorphous substitution in phyllosilicates; Variable charge under varying pH values in oxides and phyllosilicates; PZC: kaolinite = ~3.6, montmorillonite = 7.2, hematite = 5.9; Exposed -OH groups (e.g., in kaolinite) for H-bonding. | Johnson et al. 2007; Xiao et al. 2011; Zhang et al. 2014 |
| Starch-modified oxidic clays | Increased surface area from 3.98 m2/g to 8.21 m2/g after modification; Chemically stable modified product; Intact magenta properties after modification. | Reversal of surface charge from −23mV at pH = 6.5 to slightly positive values at pH = 2−9 after modification; Enhanced surface functional groups such as OH-, -COO-, and C—O after modification. | Gong et al. 2016 |
| Organoclay minerals | Decreased surface area after modification (e.g., 44 m2/g of organopalygorskite against 97 m2/g of pristine palygorskite due to pore blocking by surfactant molecules). | Reversal of surface charge from negative to positive values (e.g., −19.9 mV in pristine palygorskite against 30.6 mV after organic modification of palygorskite); Increased amino (-NH2) functional groups; Increased hydrophobicity due to long-chain alkyl group of surfactants. | Sarkar et al. 2010; Zhou et al. 2010; Sarkar et al. 2011; Sarkar et al. 2012 |
| Clay-polymer composite | Highly porous in nature; Small particle size (e.g., 100−300 µm for PDADMAC-MMT composite). | High positive charge on surface (e.g., −40.3 mV for MMT against 41.0 mV for PDADMAC-MMT at pH 7.6); Enhanced surface functional groups; Presence of hydrophobic moieties due to the inclusion of polymer. | Ray et al. 2019 |
| Magnesium aminoclays (MgAC) | Water dispersible particles; Decreased hydrodynamic diameter (e.g., 508 nm for MgAC coated nZVI against 5130 nm for bare nZVI). | Increased surface positive charge (e.g., 23.5 mV for MgAC coated nZVI against 14.5 mV for bare nZVI); Selective affinity toward hydrophobic PFAS due to enhanced -NH2 functional groups. | Arvaniti et al. 2015 |
| Clays and clay minerals | Physical properties | Chemical properties | Reference |
|---|---|---|---|
| Clay-carbon composite | High porous structure; Decreased particle size (e.g., 2.27 nm for MSW-BC-MMT composite against 17.96 nm for MMT); Increased surface area (e.g., 8.72 m2/g for MSW-BC-MMT composite against 4.33 m2/g for MSW-BC). | Enhanced hydrophobicity and functional groups (e.g., -OH, -NH2) due to carbonaceous materials (e.g., GO, biochar) and clay minerals. | Ashiq et al. 2019; Premarathna et al. 2019 |
NOTE: BC = biochar; CEC = cation exchange capacity; MgAC = magnesium aminoclay; MSW = municipal solid waste; MSW-BC = municipal solid waste biochar; nZVI = nano zero valent iron; PDAMAC = poly(diallyldimethylammonium) chloride.
SOURCE: Mukhopadhyay et al. 2021.
instantaneously (pH 3) to approximately 85 percent of PFAS (pH 11) after several hours. At pH 7 (i.e., relevant to soils), aluminum-based DWTRs also sorbed nearly 90 to 95 percent of PFAS after several hours. Zhang et al. (2021) also noted that PFAS sorption was irreversible across all pH levels. It would be important to repeat this study in soils over a broad range of pH values and other characteristics such as those where ions that compete for DWTR binding sites are present in varying concentrations.
Broadbent et al. (2025) mixed one of three different DWTRs (aluminum-, iron-, and calcium-based) into biosolids (50 g DWTR to 1 kg of anaerobically treated biosolids), then applied the mixture to a soil to raise tomatoes (i.e., considered an agricultural setting; 9 g biosolids [containing 0.45 g DWTR]/kg soil) or to a soil to raise ryegrass (i.e., considered a mine land reclamation setting; 130 g biosolids [containing 6.5 g DWTR]/kg soil). Broadbent et al. (2025) found that adding calcium-based DWTRs to biosolids reduced ryegrass PFBA concentrations. They proposed this shift could be due to the increase in soil pH from 6.5 to 7.2 caused by the DWTRs; however, the effective pKa of PFCAs in an unsaturated system given pH-dependence on surface activity is not well established (Murillo-Gelvez et al. 2023; Patel et al. 2024). Broadbent et al. (2025) also noted that PFAS plant uptake was unaffected by aluminum- or iron-based DWTRs added to biosolids, but this may have been caused by both biosolids and soils containing elevated aluminum and iron content. The latter result suggested that in order to reduce plant PFAS uptake, aluminum- and iron-based DWTRs may be better used in soils containing relatively low amounts of aluminum and iron.
Openiyi et al. (2025a) focused attention on water treatment residuals from using aluminum chlorohydrate to treat wastewater prior to re-injecting it into an aquifer or a biosolids-biochar (1 and 1.5 percent by weight, respectively) to reduce PFAS mobility in soils after biosolids land application (3 percent by weight). Biosolids-biochar showed 41 percent lower total PFAS loss from soil as compared to a control; the wastewater treatment residuals had 31-percent lower total PFAS loss as compared to a control. Both
amendments reduced PFOS losses by between 62 and 68 percent as compared to a control. The wastewater treatment residuals had higher PFAS loading than other DWTRs, but desorption of resident PFAS was low and the presence of dissolved organic matter in the solution did not affect the ability of the residuals to sorb additional PFAS (Gravesen et al. 2025). The findings of Broadbent et al. (2025) and Openiyi et al. (2025a), which are novel but limited, suggest that further research be performed to fully ascertain the use of DWTRs, singly or in combination with other amendments, for reducing PFAS phytoavailability and leaching loss from PFAS-impacted soils. A research gap could be closed by focusing efforts of combining sorbents (e.g., biochars with biochars, or biochars with DWTR) for reducing PFAS fate and transport to the greatest degree possible. This type of approach could offer a more comprehensive solution in the real world where multiple PFAS are present (Nasrollahpour et al. 2025). Additional applied research in laboratory or greenhouse trials should be performed first, followed by field investigations.
Although no conservation practice standard exists for the use of DWTRs, their use likely would fall under NRCS Nutrient Management (Code 590) conservation practice standard as DWTRs have been proven to reduce phosphorus availability. DWTRs have been proven to sorb excess phosphorus from biosolids when co-applied (Bayley et al. 2008; Ippolito et al. 2009). This approach could be taken to potentially reduce PFAS bioavailability from biosolids, though DWTRs have yet to be fully proven to sorb and reduce PFAS bioavailability. DWTRs have also been used in buffer strips to help capture phosphorus from surface water runoff (Dayton and Basta 2005; Wagner et al. 2008), and thus the potential exists to utilize DWTRs in buffer strips to reduce off-site PFAS transport. Materials similar to DWTRs have been used to capture off-site phosphorus transport in edge-of-field or in-stream containers, and this approach might prove effective at removing PFAS from waters after leaving a site. This type of approach could eventually tie into the creation of a PFAS Site Index, similar to the Phosphorus Site Index used by several U.S. states (Box 5-2). The above concepts should be further explored.
Strategies for Reducing PFAS Discharge to Surface Waters
In addition to using DWTRs to capture PFAS in buffer strips, there are other conservation practices that have been used to trap nutrients that could be experimented with to see if they would also trap PFAS. Two possibilities are reviewed: removal structures and bioreactors.
Removal Structures
To minimize PFAS-contaminated soil and water runoff as well as PFAS discharge from tile drainage, some of the approaches used to capture phosphorus from runoff and tile drainage, originated by Chad Penn at the Agricultural Research Service (ARS) of the U.S. Department of Agriculture (USDA), may be applicable with some modification for capturing PFAS. Penn and Bowen (2018) have detailed several examples of structures
that can be used to reduce phosphorus entering surface water bodies, including modular boxes, ditch filters, surface confined beds, cartridges, pond filters, blind/surface inlets, bioretention cells, and subsurface tile drain filters.1 In addition, they have detailed a suite of sorbent options and provided guidance on sorbent selection including a flow chart of things to consider. Lastly, they have included chapters for evaluating sorptive media, which is key in selecting an appropriate sorbent and designing the structure to achieve targeted goals.2 This work has been developed into a newly established conservation practice standard, Phosphorous Removal System, Code 624 (USDA 2025).
Sorbent selection and design parameters to optimize contaminant capture include sorbent capacity, sorption kinetics, estimated compound mass discharge, and desorption potential. These parameters are needed to determine the residence time and sorbent mass needed in reactors and the frequency at which sorbents will need to be replaced. Other factors to consider include sorbent availability and affordability, as well as ensuring no toxic release from the sorbent of other contaminants (e.g., heavy metals, pH, alkalinity). Additional considerations beyond the design parameters for phosphorus include what
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1 See Chapter 3 in Penn and Bowen (2018).
2 See Chapters 5 and 6 in Penn and Bowen (2018).
BOX 5-2
PFAS Site Index
Another potential future strategy to follow could be similar to the Phosphorus Site Index (P-index) approach used by several U.S. states. The P-index approach is a tool used by states to evaluate the potential for phosphorus loss from agricultural fields into waterbodies. The state of Maryland’s P-index approacha is one example of how a PFAS-index approach might be applicable. The Maryland approach evaluates potential phosphorus loss due to site transport characteristics and management practices. Site characteristics include soil test for phosphorus, soil erosion estimates, soil runoff class, leaching potential and/or subsurface drainage, distance to a water body, and priority of the receiving water body. Management characteristics include phosphorus fertilizer application rate and method or organic phosphorus application rate and method. All characteristics are scored, a phosphorus loss rating is determined, and the site of concern is placed into one of four categories regarding the potential for phosphorus movement off site. State-specific threshold models for PFAS could follow a similar approach, considering soil characteristics (e.g., via information found in the Web Soil Survey), other site characteristics, management practices, and specific PFAS present on site.
a Maryland Nutrient Management Manual. September 2000. “Section II-C Phosphorous Site Index for Maryland.” https://mda.maryland.gov/resource_conservation/pages/nm_manual.aspx.
specific PFAS need to be captured, as well as PFAS-specific concentrations deemed acceptable for final discharge.
Unlike recommended limits for phosphorus concentrations in flowing waters, which are trying to prevent eutrophication (EPA 1986), considerations of final acceptable PFAS concentrations and masses entering the nearby surface water bodies or agricultural ditch networks should be based on exposure pathways to both terrestrial and aquatic wildlife, as well as free-grazing livestock. Sorbent reactivity, kinetics, and capacity for long- versus short-chain PFAS, as well as affordability and final disposal or regeneration, will need to be considered. Collection or recovery of PFAS-sorbed materials will have an associated cost, so strategically placing sorbents in engineered structures that can be easily connected and disconnected from in-flow water sources and removed or replaced via the use of construction equipment might be a logical approach. Additionally, phosphorus-laden spent media has the potential to be used as a soil amendment in phosphorus-limiting soils or at least in soils where phosphorus levels are not problematic, but PFAS-laden spent media would be considered more in the context of a hazardous waste. Currently, the industry standard for regeneration of PFAS-laden granular activated carbon (e.g., after treating drinking water) is through thermal processes that lead to PFAS volatilization and/or destruction, depending on the temperature (DiStefano et al. 2022). While thermal approaches have been shown to return granular activated carbon to near its virgin state and with similar PFAS-sorption capacities for up to five thermal reactivation steps, literature is currently void on thermal reactivation of biochar.
Research for optimizing absorbent use for PFAS capture in runoff and tile drainage is warranted. Data resulting from some of the current bench- and field-scale efforts on the effectiveness of sorbents tilled into soil for reducing PFAS leaching and plant uptake may be useful as a starting point. However, reactor designs for PFAS in runoff and tile drain discharge will have orders of magnitude higher flow rates compared with water drainage through a soil profile, especially in the case of tile drainage. The latter will be controlled by the amount of the field being drained, diameter of the tile drains, and, of course, weather. The ultimate goal in providing guidance could be the publication of a book similar to the one published by Penn and Bowen (2018) or an updated edition with inclusion of additional chapters targeting PFAS. In addition, parallel to the P-TRAP software created to design phosphorus removal structures (Penn et al. 2021), software specific to PFAS removal could be developed.
Bioreactors
Denitrifying bioreactors have been used to reduce nutrients in runoff and tile drain discharge from agriculture fields prior to flow entering surface waters (Easton 2023). Bioreactors have been found to be more effective at nutrient removal than routing flow into constructed wetlands (Robertson and Merkley 2009). There are a variety of bioreactor designs, but all typically include organic rich materials (e.g., wood chips) to increase microbial activity and hydraulic residence time to allow sufficient time to meet target denitrification goals (Christianson et al. 2021). Guidance on construction and design of
treatment of tile drain discharge is provided by NRCS conservation practice standard Denitrifying Bioreactor (Code 605; USDA 2020). Other uses of bioreactors are being tried, but practice standards have not yet been developed.
Although efforts to increase the long-term effectiveness of these bioreactors at lower capital and maintenance costs are ongoing (Christianson and Schipper 2016; Christianson et al. 2021), there is potential that similarly designed bioreactors could be optimized to facilitate PFAS degradation. In particular, in-ditch bioreactors, similar to what is used in the construction industry to control soil runoff (referred to as wattles, which resemble large stuffed socks), may have the greatest promise due to lower costs and greater flexibility in both installation options and materials used (Christianson et al. 2017; Payne et al. 2024). These bioditches can be installed parallel to flow or perpendicular to flow, or a hybrid of these two approaches can be used. In many ways, such bioditches are similar to phosphorus (or PFAS) removal structures (discussed above), but these are designed to reduce target constituents by promoting degradation with a carbon food source versus capture via sorption mechanisms. Recent work to optimize denitrification in bioreactors includes biochar, a common sorbent material targeted for contaminant capture including PFAS (see section “Biochars” above). Biochar addition to the bioreactors was found to improve the habitat for microbial growth, enrich the denitrifying bacteria taxa, improve pH buffering, and increase nutrient retention, which all worked together to increase effective denitrification (Andras et al. 2025). Additional bioreactor designs, such as denitrifying walls, were first tested in 1996 in New Zealand (Warneke et al. 2011; Christianson et al. 2021). These walls are in essence permeable reactive barriers used in other applications to capture (when filled with sorbent) or degrade biologically or abiotically targeted contaminants to reduce off-site movement.
Microbial degradation is generally considered a favored, environmentally friendly, and typically low-cost approach to reducing pollutant loads. The biggest challenges facing the implementation of bioreactors to reduce PFAS in runoff and tile drainage are identifying and fostering a microbial community that can degrade PFAS and be maintained in the natural environment and the diversity of PFAS structures (Wackett 2022, 2025). While reduction of PFAS by microbes has been reported from laboratory-scale studies, degradation has been slow and incomplete (Huang and Jaffé 2019; Yu et al. 2020; Y. Huang et al. 2025).
Wackett (2024) posed that PFAS resistance to microbial degradation is due to limited time-of-exposure because PFAS have been manufactured for less than 100 years and microbes need more time to evolve degradation capabilities. So far, microbiologists have been unsuccessful at finding microbial evolution to degrade PFAS at military sites that have been contaminated for over five decades. Repeated exposure of PFAS to microbes in the laboratory to force the evolutionary process forward, which has been successful for some chlorinated organic compounds, has not yielded results for PFAS. There are many other reasons why microbes would not evolve well to degrade PFAS, including fluoride intercellular toxicity and energy requirements to break multiple carbon–fluorine bonds (Harris et al. 2025). Therefore, the timescale feasibility for this evolution, given there are no natural perfluorocarbons, limits any direct applicability in the near future (Wackett 2024; Walker and Chang 2024).
Addressing this challenge in the near term may be best met by combining biological processes with abiotic processes (Long et al. 2024) that used microbially generated catalysts. These catalysts lead to generation of nano-palladium particles that degrade PFOA with nitrate as the electron acceptor. Of greatest challenge among the PFAS in this regard are the perfluoroalkyl sulfonates such as PFOS, which are also some of the PFAS of greatest concern because of their high potential to bioaccumulate. Nevertheless, design and implementation of bioreactors for treatment of PFAS in runoff and tile drainage that combines both biological and abiotic processes, including enhanced retention, as well as biotically catalyzed abiotic reactions, are ripe for research, even if success is likely a long way off.
UNDERSTANDING PLANT CHARACTERISTICS THAT AFFECT PFAS UPTAKE AND ACCUMULATION
In the context of conservation practices, the selected planting of specific crops or other vegetative cover could address PFAS contamination on agricultural land via plant uptake in one of two ways: (1) by trying to minimize PFAS accumulation in harvested and grazed crops or (2) by trying to maximize plant uptake for phytoremediation. For either approach to be successful, there is an urgent need to better understand the variation in PFAS uptake among agricultural and conservation plants and of the distinct plant characteristics that influence that variation.
To start, quantification of soil-to-plant transfer of PFAS across a broader range of crops and growing conditions than currently exist in the literature is needed (Wang et al. 2020). To be most useful for crop selection decisions, these studies should be conducted in the field under real-world conditions and preferably over multiple years to capture year-to-year variability.
Plant factors already understood to influence PFAS uptake include plant physiological and biochemical characteristics. It is well established that uptake and storage of PFAS are distinctly higher in the vegetative parts of plants (leaves) than in storage organs (fruits, grains, tubers; Stahl et al. 2009; Lechner and Knapp 2011; Blaine et al. 2014; Wen et al. 2014; Wang et al. 2020; Lesmeister et al. 2021; Costello and Lee 2024). These findings are the basis of recommendations by the Maine Department of Conservation, Agriculture, and Forestry and the University of Maine Cooperative Extension for farmers impacted by PFAS (Fitzgerald et al. 2025; Maine PFAS Response Program 2025), which include switching to crops that accumulate lower levels of PFAS (e.g., growing corn and small grains instead of perennial grasses), as well as tuber crops (e.g., potato) over root crops (e.g., carrots).
Transpirational flow is seen as the primary driver of PFAS uptake and storage, with plant parts that receive larger amounts of water accumulating PFAS to a greater extent (Wang et al. 2020), although selective membranes (e.g., Casparian strip) and other transfer barriers in roots and shoots are also thought to play a role (Blaine et al. 2014; Lesmeister et al. 2021). As well, protein and lipid contents (Wen et al. 2016; Xu et al. 2021), root macrostructure (fibrous versus taproot) including root density and surface area (Lesmeister et al. 2021; Xu et al. 2021), and root exudates (Xiang et al. 2020) all have been proposed as mechanisms to explain PFAS uptake difference among plant
species and cultivars. Further research is needed to investigate the relative importance of these factors.
Differences in PFAS transfer factors among crop species also have been linked to transpiration differences related to the length and seasonality of the vegetative period and the crop’s transpirational capacity (Blaine et al. 2014; Ghisi et al. 2019). For instance, in their field study at a highly contaminated site, Liu et al. (2017) observed more than 11-fold higher PFAS concentrations (a sum of 12) in wheat rather than corn; they attributed this difference to differences in the plants’ transpirational coefficients (450–600 g water/g dry weight versus 250–300 g water/g dry weight, respectively).
Differences in transpiration rates is one of the hypotheses presented by Simones et al. (2024) to explain their observation over 2 years of higher PFOS uptake by a grass-based perennial forage in aftermath growth harvested in August or September (“second cut” harvested in August or September) than in the initial growth for the season (“first cut” harvested in June). Similar observations were reported by Stahl et al. (2009) and Openiyi et al. (2025b) in greenhouse studies with perennial ryegrass and forage grasses, respectively. Another possible explanation provided by Simones et al. (2024) for this seasonality effect was that root and above-ground structure may have been different between the two growing periods. Perennial forage regrowth is known to have a higher leaf-to-stem/pseudostem ratio (Ball et al. 2001), and leafier plants may exhibit increased PFAS accumulation (Nassazzi et al. 2025) as discussed above. Clearly, more research is warranted.
There is also currently a complete lack of research in PFAS translocation mechanisms from soil into tree fruiting bodies. Gobelius et al. (2017) studied PFAS accumulation in tree species (all nonfruit-bearing) near a firefighting training facility that used aqueous film-forming foams for around 20 years. The relative accumulation of PFAS in tree plant components mirrored that found in annual plant components; birch leaves contained the greatest amount of PFAS (12–97 ng/g wet weight), followed by twigs (5–40 ng/g wet weight), trunk (1–19 ng/g wet weight), and roots (3–6 ng/g wet weight). Spruce tree components followed a somewhat similar order. PFAS (particularly long-chain PFAS) have been shown to bind to proteins as was observed for PFOA and PFOS, which favorably deposited in protein-rich versus lipid-rich root tissues (Wen et al. 2016). However, this concept is unlikely to translate to preferential accumulation in the nuts of nut-bearing trees based on what has been observed for corn kernel and plant seeds. Plant physiology controls whether PFAS in the transpiration stream (xylem) will off-load into the phloem, which appears to be limited. For example, Lazo and Lee (2024) observed no accumulation of PFOS in the bean of six different soybean varieties while PFOS concentrations were high in the other tissues: leaves containing more than stems and stems more than pods. As discussed in Chapter 2, short-chain PFCAs structurally mimic fatty acids, whereas PFOS does not. Therefore, differences in transport from the transpiration stream into the phloem and subsequently into seeds and fruits are likely due to non-specific fatty acid transporters.
Wang et al. (2020) called for increasing the level of genetic understanding of PFAS uptake into various crops. This latter point is currently being tackled by USDA–ARS, which is screening for differences in crop varieties that lead to reduced PFAS uptake into
edible plant parts, followed by genetic testing.3 In the area of phytoremediation, results have been mixed, with rates of PFAS uptake having shown that this approach may have promise for some shorter-chain PFAS but is not a viable option for longer-chain legacy compounds (Nason et al. 2024). Researchers are looking into microbe-assisted, chelate-assisted, and genetic-engineering approaches (Naveed et al. 2024), as well as nanotechnology-assisted techniques to enhance plant uptake and phytoremediation efficacy (C. Huang et al. 2025).
Research is also needed to determine and use appropriate vegetative covers that are not detrimental to the health of wildlife when consumed. If uptake of PFAS by vegetative conservation covers causes harmful impacts to wildlife (or grazing livestock), the result would be at odds with conservation practices such as Upland Wildlife Habitat Management (Code 645).
Finally, how crop management affects plant PFAS uptake is a topic that has barely been addressed but is of great importance in the context of conservation practices. Researchers have investigated fertilization (Adu et al. 2025) and intercropping (Scearce et al. 2025) on crop PFAS uptake with mixed effects. No known studies exist on how PFAS uptake is influenced by conservation tillage/no-till, which is known to affect root distribution under certain conditions (Ruis and Blanco-Canqui 2024), crop rotation, crop density, or irrigation. These are major research gaps.
PFAS MITIGATION IN LIVESTOCK
Mitigating PFAS in livestock that are exposed to PFAS-contaminated water or feed is a significant challenge for which research is needed, although some practical guidance has been developed. Animal product consumption tends to be the largest human exposure pathway for PFOS and PFOA (EFSA Panel on Contaminant in the Food Chain et al. 2020). Dairy is of particular concern because whole-plant forages, which comprise a major portion of dairy livestock diets, have relatively high rates of PFAS uptake and therefore PFAS bioaccumulates and biomagnifies in animals and their milk (Hossini et al. 2025). Representative practical guidance from approaches implemented in Maine include the following (Maine PFAS Response Program 2025):
- Take advantage of the fact that PFAS levels in animals will decrease over time once the source of PFAS in the diet is eliminated. The half-life of PFOS in milk and beef tissues is estimated to be between 8 and 12 weeks (Astmann et al. 2025). Switching beef animals to unaffected pastures or “clean feed” during the finishing stage can reduce PFAS levels in subsequent food products.
- Dilute affected feed with “clean feed” to lower PFAS intake for livestock. For hay, mark and keep records of which hay bales come from which fields. For silage, silos or bunkers must be segregated by fields or lots should be adequately mixed during fillings to avoid PFAS “hot spots” when feeding out.
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3 See https://www.ars.usda.gov/pacific-west-area/riverside-ca/agricultural-water-efficiency-and-salinity-research-unit/research/mitigation-and-remediation-of-pfas.
- Use soil tests, not forage tests, to assess the potential risk of milk or meat contamination. In the soil–to–forage–to–animal pathway, PFAS concentrations are the lowest in forage, so it is possible to have detectable milk and meat concentrations even if forage is non-detect.
Although guidance is emerging, more research is needed to support continued development of management practices. Livestock studies have focused on a very limited subset of perfluoroalkyl acids (PFAAs), primarily PFOS and PFOA, with an even smaller number including the PFCAs perfluorononanoic acid (PFNA), perfluorodecanoic acid (PFDA), perfluoroundecanoic acid (PFUnA), and perfluorododecanoic acid (PFDoDA) and the PFSAs perfluorobutanesulfonic acid (PFBS), perfluoropentanesulfonic acid (PFPeS), and perfluoroheptanesulfonic acid (PFHpS) (Lupton et al. 2012; Kowalcyzk et al. 2013; Vestergren et al. 2013; Drew et al. 2021, 2022; Chou et al. 2023; Mikkonen et al. 2023). PFAA elimination from livestock occurs via lactation, urine, and feces, with renal excretion generally considered the primary route of elimination (Mikkonen et al. 2023). For studied PFAS, elimination half-lives in serum increase with increasing chain length (e.g., 9.4, 46.0, and 67.3 days for perfluorohexane sulfonate [PFHxS], PFHpS, and PFOS, respectively) (Drew et al. 2022). Physiologically based pharmacokinetic models have been used to estimate the time for concentrations to decrease to levels below risk-based consumption limits—known as withdrawal intervals—for PFOA, PFHxS, and PFOS in milk (Chou et al. 2023). Based on the affiliated web interface, withdrawal intervals for PFOA, PFHxS, and PFOS following 2 years of exposure to 1 µg/L were <1 day, <1 day, and around 1,875 days, respectively; however, models such as these should be updated to reflect more recent regulatory limits and newly regulated PFAS and validated in real-world scenarios. Research focused on uptake and especially elimination will inform how conservation practices (e.g., Feed Management, Code 592) may be used to mitigate PFAS impacts in livestock.
Research is needed to explore opportunities to interrupt PFAS from cycling on farms and in agrifood systems through animal waste management. Many of the technologies and approaches being explored to separate and destroy PFAS in biosolids could be applied to animal manures, such as foam fractionation (Smith et al. 2023), anaerobic digestion (Li et al. 2021), and mechanochemical degradation, pyrolysis/gasification, and supercritical water oxidation (Berg et al. 2022). If promising, these approaches could be incorporated into or coupled with existing conservation practices that target manure management (e.g., Anaerobic Digestion [Code 366]; Waste Storage Facility [Code 313]; Waste Separation Facility [Code 632]; and Waste Treatment [Code 629]).
Conclusion 5-1: Applied research that advances understanding of PFAS fate and transport in different types of soils, develops better mechanisms by which to trap or sequester PFAS, and minimizes PFAS uptake in plants and animals could improve the ability of conservation practices to address PFAS contamination on agricultural land.
Conclusion 5-2: A coordinated, national network of researchers focused on the identified areas of applied research would help close information gaps and provide practical knowledge for managing PFAS contamination in U.S. agricultural systems.
Conclusion 5-3: The results of such research and coordination could be used to continually improve existing resources and provide needed resources identified in the suggested framework to advance the ability of the FPAC agencies to respond to the impacts of PFAS contamination on agricultural land.
The issues associated with PFAS in the environment are many. Means by which PFAS contamination can be mitigated still need further research. This chapter took into account the committee’s expertise both within and outside the PFAS space, utilizing combined experiences and education to derive potential applied research paths that may be of focus in the future. The described paths include dealing with PFAS in situ (i.e., in soils) by potentially reducing PFAS bioavailability, reducing PFAS from entering waters or reducing PFAS presence within water sources, understanding the complex nature of PFAS within plants and potentially altering cropping systems accordingly, and mitigating PFAS in animals, especially livestock. As the general public’s awareness of PFAS is heightened, it will be increasing important to face the challenges in communicating research results in a way that educates but does not exacerbate unfounded concerns or fears (Preisendanz et al. 2025).
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