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==PFAS Ex Situ Water Treatment==
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==Remediation of Stormwater Runoff Contaminated by Munition Constituents==  
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Well-developed ''ex situ'' treatment technologies applicable to treatment of [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | perfluoroalkyl and polyfluoroalkyl substances (PFAS)]] in drinking water and non-potable groundwater include membrane filtration (reverse osmosis and nanofiltration), activated carbon adsorption (granular and powdered), and anion exchange.  There are also a variety of separation and destructive technologies in various stages of development.  Some of these processes may also be applicable to more complex matrices including wastewater and landfill leachate.   
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Past and ongoing military operations have resulted in contamination of surface soil with [[Munitions Constituents | munition constituents (MC)]], which have human and environmental health impacts.  These compounds can be transported off site via stormwater runoff during precipitation events.  Technologies to “trap and treat” surface runoff before it enters downstream receiving bodies (e.g., streams, rivers, ponds) (see Figure 1), and which are compatible with ongoing range activities are needed.  This article describes a passive and sustainable approach for effective management of munition constituents in stormwater runoff.   
 
<div style="float:right;margin:0 0 2em 2em;">__TOC__</div>
 
<div style="float:right;margin:0 0 2em 2em;">__TOC__</div>
  
 
'''Related Article(s):'''
 
'''Related Article(s):'''
  
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* [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS)]]
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*[[Munitions Constituents]]
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* [[PFAS Transport and Fate]]
 
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* [[PFAS Sources]]
 
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* [[PFAS Soil Remediation Technologies]]
 
  
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'''Contributor(s):''' [[Dr. Scott Grieco]] and [[James Hatton]]
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'''Contributor:''' Mark E. Fuller
  
 
'''Key Resource(s):'''
 
'''Key Resource(s):'''
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*SERDP Project ER19-1106: Development of Innovative Passive and Sustainable Treatment Technologies for Energetic Compounds in Surface Runoff on Active Ranges
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==Background==
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===Surface Runoff Characteristics and Treatment Approaches===
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[[File: FullerFig1.png | thumb | 400 px | Figure 1. Conceptual model of passive trap and treat approach for MC removal from stormwater runoff]]
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During large precipitation events the rate of water deposition exceeds the rate of water infiltration, resulting in surface runoff (also called stormwater runoff). Surface characteristics including soil texture, presence of impermeable surfaces (natural and artificial), slope, and density and type of vegetation all influence the amount of surface runoff from a given land area. The use of passive systems such as retention ponds and biofiltration cells for treatment of surface runoff is well established for urban and roadway runoff. Treatment in those cases is typically achieved by directing runoff into and through a small constructed wetland, often at the outlet of a retention basin, or via filtration by directing runoff through a more highly engineered channel or vault containing the treatment materials. Filtration based technologies have proven to be effective for the removal of metals, organics, and suspended solids<ref>Sansalone, J.J., 1999. In-situ performance of a passive treatment system for metal source control. Water Science and Technology, 39(2), pp. 193-200. [https://doi.org/10.1016/S0273-1223(99)00023-2 doi: 10.1016/S0273-1223(99)00023-2]</ref><ref>Deletic, A., Fletcher, T.D., 2006. Performance of grass filters used for stormwater treatment—A field and modelling study. Journal of Hydrology, 317(3-4), pp. 261-275. [http://dx.doi.org/10.1016/j.jhydrol.2005.05.021 doi: 10.1016/j.jhydrol.2005.05.021]</ref><ref>Grebel, J.E., Charbonnet, J.A., Sedlak, D.L., 2016. Oxidation of organic contaminants by manganese oxide geomedia for passive urban stormwater treatment systems. Water Research, 88, pp. 481-491. [http://dx.doi.org/10.1016/j.watres.2015.10.019 doi: 10.1016/j.watres.2015.10.019]</ref><ref>Seelsaen, N., McLaughlan, R., Moore, S., Ball, J.E., Stuetz, R.M., 2006. Pollutant removal efficiency of alternative filtration media in stormwater treatment. Water Science and Technology, 54(6-7), pp. 299-305. [https://doi.org/10.2166/wst.2006.617 doi: 10.2166/wst.2006.617]</ref>.
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===Surface Runoff on Ranges===
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[[File: FullerFig2.png | thumb | 500 px | Figure 2. Conceptual illustration of munition constituent production and transport on military ranges. Mesoscale residues are qualitatively defined as being easily visible to the naked eye (e.g., from around 50 µm to multiple cm in size) and less likely to be transported by moving water.  Microscale residues are defined as <50 µm down to below 1 µm, and more likely to be entrained in, and transported by, moving water as particulates. Blue arrows represent possible water flow paths and include both dissolved and solid phase energetics. The red vertical arrow represents the predominant energetics dissolution process in close proximity to the residues due to precipitation.]]
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Surface runoff represents a major potential mechanism through which energetics residues and related materials are transported off site from range soils to groundwater and surface water receptors (Figure 2). This process is particularly important for energetics that are water soluble (e.g., [[Wikipedia: Nitrotriazolone | NTO]] and [[Wikipedia: Nitroguanidine | NQ]]) or generate soluble daughter products (e.g., [[Wikipedia: 2,4-Dinitroanisole | DNAN]] and [[Wikipedia: TNT | TNT]]). While traditional MC such as [[Wikipedia: RDX | RDX]] and [[Wikipedia: HMX | HMX]] have limited aqueous solubility, they also exhibit recalcitrance to degrade under most natural conditions. RDX and [[Wikipedia: Perchlorate | perchlorate]] are frequent groundwater contaminants on military training ranges. While actual field measurements of energetics in surface runoff are limited, laboratory experiments have been performed to predict mobile energetics contamination levels based on soil mass loadings<ref>Cubello, F., Polyakov, V., Meding, S.M., Kadoya, W., Beal, S., Dontsova, K., 2024. Movement of TNT and RDX from composition B detonation residues in solution and sediment during runoff. Chemosphere, 350, Article 141023. [https://doi.org/10.1016/j.chemosphere.2023.141023 doi: 10.1016/j.chemosphere.2023.141023]</ref><ref>Karls, B., Meding, S.M., Li, L., Polyakov, V., Kadoya, W., Beal, S., Dontsova, K., 2023. A laboratory rill study of IMX-104 transport in overland flow. Chemosphere, 310, Article 136866. [https://doi.org/10.1016/j.chemosphere.2022.136866 doi: 10.1016/j.chemosphere.2022.136866]&nbsp; [[Media: KarlsEtAl2023.pdf | Open Access Article]]</ref><ref>Polyakov, V., Beal, S., Meding, S.M., Dontsova, K., 2025. Effect of gypsum on transport of IMX-104 constituents in overland flow under simulated rainfall. Journal of Environmental Quality, 54(1), pp. 191-203. [https://doi.org/10.1002/jeq2.20652 doi: 10.1002/jeq2.20652]&nbsp; [[Media: PolyakovEtAl2025.pdf | Open Access Article.pdf]]</ref><ref>Polyakov, V., Kadoya, W., Beal, S., Morehead, H., Hunt, E., Cubello, F., Meding, S.M., Dontsova, K., 2023. Transport of insensitive munitions constituents, NTO, DNAN, RDX, and HMX in runoff and sediment under simulated rainfall. Science of the Total Environment, 866, Article 161434. [https://doi.org/10.1016/j.scitotenv.2023.161434 doi: 10.1016/j.scitotenv.2023.161434]&nbsp; [[Media: PolyakovEtAl2023.pdf | Open Access Article.pdf]]</ref><ref>Price, R.A., Bourne, M., Price, C.L., Lindsay, J., Cole, J., 2011. Transport of RDX and TNT from Composition-B Explosive During Simulated Rainfall. In: Environmental Chemistry of Explosives and Propellant Compounds in Soils and Marine Systems: Distributed Source Characterization and Remedial Technologies. American Chemical Society, pp. 229-240. [https://doi.org/10.1021/bk-2011-1069.ch013 doi: 10.1021/bk-2011-1069.ch013]</ref>. For example, in a previous small study, MC were detected in surface runoff from an active live-fire range<ref>Fuller, M.E., 2015. Fate and Transport of Colloidal Energetic Residues. Department of Defense Strategic Environmental Research and Development Program (SERDP), Project ER-1689. [https://serdp-estcp.mil/projects/details/10760fd6-fb55-4515-a629-f93c555a92f0 Project Website]&nbsp;&nbsp; [[Media: ER-1689-FR.pdf | Final Report.pdf]]</ref>, and more recent sampling has detected MC in marsh surface water adjacent to the same installation (personal communication).  Another recent report from Canada also detected RDX in both surface runoff and surface water at low part per billion levels in a survey of several military demolition sites<ref>Lapointe, M.-C., Martel, R., Diaz, E., 2017. A Conceptual Model of Fate and Transport Processes for RDX Deposited to Surface Soils of North American Active Demolition Sites. Journal of Environmental Quality, 46(6), pp. 1444-1454. [https://doi.org/10.2134/jeq2017.02.0069 doi: 10.2134/jeq2017.02.0069]</ref>. However, overall, data regarding the MC contaminant profile of surface runoff from ranges is very limited, and the possible presence of non-energetic constituents (e.g., metals, binders, plasticizers) in runoff has not been examined.  Additionally, while energetics-contaminated surface runoff is an important concern, mitigation technologies specifically for surface runoff  have not yet been developed and widely deployed in the field.  To effectively capture and degrade MC and associated compounds that are present in surface runoff, novel treatment media are needed to sorb a broad range of energetic materials and to transform the retained compounds through abiotic and/or microbial processes.
  
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*[https://www.waterrf.org/resource/treatment-mitigation-strategies-poly-and-perfluorinated-chemicals Water Research Foundation (Drinking Water): Treatment Mitigation Strategies for PFAS]<ref name="Dickenson2016">Dickenson, E. and Higgins, C., 2016. Treatment Mitigation Strategies for Poly- and Perfluoroalkyl Substances, Report Number 4322. Water Research Foundation, Denver, Colorado. 123 pages. ISBN 978-1-60573-234-3</ref>
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Surface runoff of organic and inorganic contaminants from live-fire ranges is a challenging issue for the Department of Defense (DoD). Potentially even more problematic is the fact that inputs to surface waters from large testing and training ranges typically originate from multiple sources, often encompassing hundreds of acres. No available technologies are currently considered effective for controlling non-point source energetics-laden surface runoff. While numerous technologies exist to treat collected explosives residues, contaminated soil and even groundwater, the decentralized nature and sheer volume of military range runoff have precluded the use of treatment technologies at full scale in the field.
  
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*[https://pfas-1.itrcweb.org/12-treatment-technologies/#12_2 Interstate Technical and Regulatory Council: PFAS Liquids Treatment Technologies]<ref name="ITRC2020">Interstate Technology and Regulatory Council (ITRC), 2020. PFAS Technical and Regulatory Guidance Document and Fact Sheets, PFAS-1. PFAS Team, Washington, DC.  [https://pfas-1.itrcweb.org/ Website]&nbsp;&nbsp; [[Media: ITRC_PFAS-1.pdf | Report.pdf]]</ref> 
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==Range Runoff Treatment Technology Components==
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Based on the conceptual foundation of previous research into surface water runoff treatment for other contaminants, with a goal to “trap and treat” the target compounds, the following components were selected for inclusion in the technology developed to address range runoff contaminated with energetic compounds.
  
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*[https://www.sciencedirect.com/science/article/pii/S0301479717307934 Novel treatment technologies for PFAS compounds: A critical review.]<ref name="Kucharzyk2017"> Kucharzyk, K.H., Darlington, R., Benotti, M., Deeb, R. and Hawley, E., 2017. Novel treatment technologies for PFAS compounds: A critical review. Journal of Environmental Management, 204(2), pp. 757-764. [https://doi.org/10.1016/j.jenvman.2017.08.016 DOI: 10.1016/j.jenvman.2017.08.016]&nbsp;&nbsp; Manuscript available from: [https://www.researchgate.net/profile/Katarzyna_kate_Kucharzyk/publication/319125507_Novel_treatment_technologies_for_PFAS_compounds_A_critical_review/links/5a06590b4585157013a3be77/Novel-treatment-technologies-for-PFAS-compounds-A-critical-review.pdf ResearchGate].</ref>
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===Peat===
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Previous research demonstrated that a peat-based system provided a natural and sustainable sorptive medium for organic explosives such as HMX, RDX, and TNT, allowing much longer residence times than predicted from hydraulic loading alone<ref>Fuller, M.E., Hatzinger, P.B., Rungkamol, D., Schuster, R.L., Steffan, R.J., 2004. Enhancing the attenuation of explosives in surface soils at military facilities: Combined sorption and biodegradation. Environmental Toxicology and Chemistry, 23(2), pp. 313-324. [https://doi.org/10.1897/03-187 doi: 10.1897/03-187]</ref><ref>Fuller, M.E., Lowey, J.M., Schaefer, C.E., Steffan, R.J., 2005. A Peat Moss-Based Technology for Mitigating Residues of the Explosives TNT, RDX, and HMX in Soil. Soil and Sediment Contamination: An International Journal, 14(4), pp. 373-385. [https://doi.org/10.1080/15320380590954097 doi: 10.1080/15320380590954097]</ref><ref name="FullerEtAl2009">Fuller, M.E., Schaefer, C.E., Steffan, R.J., 2009. Evaluation of a peat moss plus soybean oil (PMSO) technology for reducing explosive residue transport to groundwater at military training ranges under field conditions. Chemosphere, 77(8), pp. 1076-1083. [https://doi.org/10.1016/j.chemosphere.2009.08.044 doi: 10.1016/j.chemosphere.2009.08.044]</ref><ref>Hatzinger, P.B., Fuller, M.E., Rungkamol, D., Schuster, R.L., Steffan, R.J., 2004. Enhancing the attenuation of explosives in surface soils at military facilities: Sorption-desorption isotherms. Environmental Toxicology and Chemistry, 23(2), pp. 306-312. [https://doi.org/10.1897/03-186 doi: 10.1897/03-186]</ref><ref name="SchaeferEtAl2005">Schaefer, C.E., Fuller, M.E., Lowey, J.M., Steffan, R.J., 2005. Use of Peat Moss Amended with Soybean Oil for Mitigation of Dissolved Explosive Compounds Leaching into the Subsurface: Insight into Mass Transfer Mechanisms. Environmental Engineering Science, 22(3), pp. 337-349. [https://doi.org/10.1089/ees.2005.22.337 doi: 10.1089/ees.2005.22.337]</ref>. Peat moss represents a bioactive environment for treatment of the target contaminants. While the majority of the microbial reactions are aerobic due to the presence of measurable dissolved oxygen in the bulk solution, anaerobic reactions (including methanogenesis) can occur in microsites within the peat. The peat-based substrate acts not only as a long term electron donor as it degrades but also acts as a strong sorbent. This is important in intermittently loaded systems in which a large initial pulse of MC can be temporarily retarded on the peat matrix and then slowly degraded as they desorb<ref name="FullerEtAl2009"/><ref name="SchaeferEtAl2005"/>. This increased residence time enhances the biotransformation of energetics and promotes the immobilization and further degradation of breakdown products. Abiotic degradation reactions are also likely enhanced by association with the organic-rich peat (e.g., via electron shuttling reactions of [[Wikipedia: Humic substance | humics]])<ref>Roden, E.E., Kappler, A., Bauer, I., Jiang, J., Paul, A., Stoesser, R., Konishi, H., Xu, H., 2010. Extracellular electron transfer through microbial reduction of solid-phase humic substances. Nature Geoscience, 3, pp. 417-421. [https://doi.org/10.1038/ngeo870 doi: 10.1038/ngeo870]</ref>.
  
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*[https://www.liebertpub.com/doi/abs/10.1089/ees.2016.0233 Degradation and removal methods for perfluoroalkyl and polyfluoroalkyl substances in water]<ref name="Merino2016">Merino, N., Qu, Y., Deeb, R.A., Hawley, E.L., Hoffmann, M.R., and Mahendra, S., 2016. Degradation and Removal Methods for Perfluoroalkyl and Polyfluoroalkyl Substances in Water. Environmental Engineering Science, 33(9), pp. 615-649. [https://doi.org/10.1089/ees.2016.0233 DOI: 10.1089/ees.2016.0233]</ref>
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===Soybean Oil===
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Modeling has indicated that peat moss amended with crude soybean oil would significantly reduce the flux of dissolved TNT, RDX, and HMX through the vadose zone to groundwater compared to a non-treated soil (see [https://serdp-estcp.mil/projects/details/20e2f05c-fd50-4fd3-8451-ba73300c7531 ESTCP ER-200434]). The technology was validated in field soil plots, showing a greater than 500-fold reduction in the flux of dissolved RDX from macroscale Composition B detonation residues compared to a non-treated control plot<ref name="FullerEtAl2009"/>. Laboratory testing and modeling indicated that the addition of soybean oil increased the biotransformation rates of RDX and HMX at least 10-fold compared to rates observed with peat moss alone<ref name="SchaeferEtAl2005"/>. Subsequent experiments also demonstrated the effectiveness of the amended peat moss material for stimulating perchlorate transformation when added to a highly contaminated soil (Fuller et al., unpublished data). These previous findings clearly demonstrate the effectiveness of peat-based materials for mitigating transport of both organic and inorganic energetic compounds through soil to groundwater.  
  
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==Established PFAS Treatment Technologies==
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===Biochar===
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Three technologies are well demonstrated for removal of [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | PFAS]] from drinking water and non-potable groundwater (as described below):
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Recent reports have highlighted additional materials that, either alone, or in combination with electron donors such as peat moss and soybean oil, may further enhance the sorption and degradation of surface runoff contaminants, including both legacy energetics and [[Wikipedia: Insensitive_munition#Insensitive_high_explosives | insensitive high explosives (IHE)]].  For instance, [[Wikipedia: Biochar | biochar]], a type of black carbon, has been shown to not only sorb a wide range of organic and inorganic contaminants including MCs<ref>Ahmad, M., Rajapaksha, A.U., Lim, J.E., Zhang, M., Bolan, N., Mohan, D., Vithanage, M., Lee, S.S., Ok, Y.S., 2014. Biochar as a sorbent for contaminant management in soil and water: A review. Chemosphere, 99, pp. 19-33. [https://doi.org/10.1016/j.chemosphere.2013.10.071 doi: 10.1016/j.chemosphere.2013.10.071]</ref><ref>Mohan, D., Sarswat, A., Ok, Y.S., Pittman, C.U., 2014. Organic and inorganic contaminants removal from water with biochar, a renewable, low cost and sustainable adsorbent – A critical review. Bioresource Technology, 160, pp. 191-202. [https://doi.org/10.1016/j.biortech.2014.01.120 doi: 10.1016/j.biortech.2014.01.120]</ref><ref>Oh, S.-Y., Seo, Y.-D., Jeong, T.-Y., Kim, S.-D., 2018. Sorption of Nitro Explosives to Polymer/Biomass-Derived Biochar. Journal of Environmental Quality, 47(2), pp. 353-360. [https://doi.org/10.2134/jeq2017.09.0357 doi: 10.2134/jeq2017.09.0357]</ref><ref>Xie, T., Reddy, K.R., Wang, C., Yargicoglu, E., Spokas, K., 2015. Characteristics and Applications of Biochar for Environmental Remediation: A Review. Critical Reviews in Environmental Science and Technology, 45(9), pp. 939-969. [https://doi.org/10.1080/10643389.2014.924180 doi: 10.1080/10643389.2014.924180]</ref>, but also to facilitate their degradation<ref>Oh, S.-Y., Cha, D.K., Kim, B.-J., Chiu, P.C., 2002. Effect of adsorption to elemental iron on the transformation of 2,4,6-trinitrotoluene and hexahydro-1,3,5-trinitro-1,3,5-triazine in solution. Environmental Toxicology and Chemistry, 21(7), pp. 1384-1389. [https://doi.org/10.1002/etc.5620210708 doi: 10.1002/etc.5620210708]</ref><ref>Ye, J., Chiu, P.C., 2006. Transport of Atomic Hydrogen through Graphite and its Reaction with Azoaromatic Compounds. Environmental Science and Technology, 40(12), pp. 3959-3964. [https://doi.org/10.1021/es060038x doi: 10.1021/es060038x]</ref><ref name="OhChiu2009">Oh, S.-Y., Chiu, P.C., 2009. Graphite- and Soot-Mediated Reduction of 2,4-Dinitrotoluene and Hexahydro-1,3,5-trinitro-1,3,5-triazine. Environmental Science and Technology, 43(18), pp. 6983-6988. [https://doi.org/10.1021/es901433m doi: 10.1021/es901433m]</ref><ref name="OhEtAl2013">Oh, S.-Y., Son, J.-G., Chiu, P.C., 2013. Biochar-mediated reductive transformation of nitro herbicides and explosives. Environmental Toxicology and Chemistry, 32(3), pp. 501-508. [https://doi.org/10.1002/etc.2087 doi: 10.1002/etc.2087]&nbsp;&nbsp; [[Media: OhEtAl2013.pdf | Open Access Article.pdf]]</ref><ref name="XuEtAl2010">Xu, W., Dana, K.E., Mitch, W.A., 2010. Black Carbon-Mediated Destruction of Nitroglycerin and RDX by Hydrogen Sulfide. Environmental Science and Technology, 44(16), pp. 6409-6415. [https://doi.org/10.1021/es101307n doi: 10.1021/es101307n]</ref><ref>Xu, W., Pignatello, J.J., Mitch, W.A., 2013. Role of Black Carbon Electrical Conductivity in Mediating Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) Transformation on Carbon Surfaces by Sulfides. Environmental Science and Technology, 47(13), pp. 7129-7136. [https://doi.org/10.1021/es4012367 doi: 10.1021/es4012367]</ref>. Depending on the source biomass and [[Wikipedia: Pyrolysis| pyrolysis]] conditions, biochar can possess a high [[Wikipedia: Specific surface area | specific surface area]] (on the order of several hundred m<small><sup>2</sup></small>/g)<ref>Zhang, J., You, C., 2013. Water Holding Capacity and Absorption Properties of Wood Chars. Energy and Fuels, 27(5), pp. 2643-2648. [https://doi.org/10.1021/ef4000769 doi: 10.1021/ef4000769]</ref><ref>Gray, M., Johnson, M.G., Dragila, M.I., Kleber, M., 2014. Water uptake in biochars: The roles of porosity and hydrophobicity. Biomass and Bioenergy, 61, pp. 196-205. [https://doi.org/10.1016/j.biombioe.2013.12.010 doi: 10.1016/j.biombioe.2013.12.010]</ref> and hence a high sorption capacity.  Biochar and other black carbon also exhibit especially high affinity for [[Wikipedia: Nitro compound | nitroaromatic compounds (NACs)]] including TNT and 2,4-dinitrotoluene (DNT)<ref>Sander, M., Pignatello, J.J., 2005. Characterization of Charcoal Adsorption Sites for Aromatic Compounds:  Insights Drawn from Single-Solute and Bi-Solute Competitive Experiments. Environmental Science and Technology, 39(6), pp. 1606-1615. [https://doi.org/10.1021/es049135l doi: 10.1021/es049135l]</ref><ref name="ZhuEtAl2005">Zhu, D., Kwon, S., Pignatello, J.J., 2005. Adsorption of Single-Ring Organic Compounds to Wood Charcoals Prepared Under Different Thermochemical Conditions. Environmental Science and Technology 39(11), pp. 3990-3998. [https://doi.org/10.1021/es050129e doi: 10.1021/es050129e]</ref><ref name="ZhuPignatello2005">Zhu, D., Pignatello, J.J., 2005. Characterization of Aromatic Compound Sorptive Interactions with Black Carbon (Charcoal) Assisted by Graphite as a Model. Environmental Science and Technology, 39(7), pp. 2033-2041. [https://doi.org/10.1021/es0491376 doi: 10.1021/es0491376]</ref>. This is due to the strong [[Wikipedia: Pi-interaction | ''&pi;-&pi;'' electron donor-acceptor interactions]] between electron-rich graphitic domains in black carbon and the electron-deficient aromatic ring of the NAC
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* membrane filtration including [[wikipedia: Reverse osmosis | reverse osmosis (RO)]] and [[Wikipedia: Nanofiltration | nanofiltration (NF)]]  
 
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* granular [[Wikipedia: Activated carbon | activated carbon]] (GAC) and powdered activated carbon (PAC) adsorption
 
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* [[wikipedia: Ion_exchange | anion exchange (IX)]]
 
  
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However, these technologies are less demonstrated for removal of PFAS from more complex matrices such as wastewater and leachate.  
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==PFAS Screening Levels for Human Health and Ecological Risk Assessments==
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Site-specific considerations that affect the selection of optimum treatment technologies for a given site include water chemistry, required flow rate, treatment criteria, waste residual generation, residual disposal options, and operational complexity.  Treatability studies with site water are highly recommended because every site has different factors that may affect engineering design for these technologies.
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===Human Health Screening Levels===
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Human health screening levels for PFAS have been modified multiple times over the last decade and, in the United States, are currently available for drinking water and soil exposures as Maximum Contaminant Levels (MCLs) and USEPA Regional Screening Levels (RSLs). USEPA finalized a National Primary Drinking Water Regulation (NPDWR) for six PFAS<ref name="USEPA2024b"/>:
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*Perfluorooctanoic acid (PFOA)
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*Perfluorooctane sulfonic acid (PFOS)
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*Perfluorohexane sulfonic acid (PFHxS)
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*Perfluorononanoic acid (PFNA)
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*Hexafluoropropylene oxide dimer acid (HFPO-DA, commonly known as GenX chemicals)
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*Perfluorobutane sulfonic acid (PFBS)
  
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===Membrane Filtration===
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MCLs are enforceable drinking water standards based on the most recently available toxicity information that consider available treatment technologies and costs. The MCLs for PFAS include a Hazard Index of 1 for combined exposures to four PFAS. RSLs are developed for use in risk assessments and include soil and tap water screening levels for multiple PFAS. Soil RSLs are based on residential/unrestricted and commercial/industrial land uses, and calculations of site-specific RSLs are available.   
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[[File: revOsmosisPlant.png | thumb | 500px | Figure 1.  A RO municipal drinking water plant in Arizona]]
 
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Given their ability to remove dissolved contaminants at a molecular size level, RO and some NF membranes can be highly effective for PFAS removal. For RO systems (Figure 1), several studies have demonstrated effective removal of perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) (see [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | PFAS]] for nomenclature) from drinking water with removal rates well above 90%<ref name="Tang2006">Tang, C.Y., Fu, Q.S., Robertson, A.P., Criddle, C.S., and Leckie, J.O., 2006. Use of Reverse Osmosis Membranes to Remove Perfluorooctane Sulfonate (PFOS) from Semiconductor Wastewater. Environmental Science and Technology, 40(23), pp. 7343-7349.  [https://doi.org/10.1021/es060831q DOI: 10.1021/es060831q]</ref><ref name="Flores2013">Flores, C., Ventura, F., Martin-Alonso, J., and Caixach, J., 2013. Occurrence of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) in NE Spanish surface waters and their removal in a drinking water treatment plant that combines conventional and advanced treatments in parallel lines. Science of the Total environment, 461, 618-626.  [https://doi.org/10.1016/j.scitotenv.2013.05.026 DOI: 10.1016/j.scitotenv.2013.05.026]</ref><ref name="Appleman2014">Appleman, T.D., Higgins, C.P., Quiñones, O., Vanderford, B.J., Kolstad, C., Zeigler-Holady, J.C., and Dickenson, E.R., 2014. Treatment of poly- and perfluoroalkyl substances in US full-scale water treatment systems. Water Research, 51, pp. 246-255.  [https://doi.org/10.1016/j.watres.2013.10.067 DOI: 10.1016/j.watres.2013.10.067]</ref>. RO potable water reuse treatment systems implemented in California have also demonstrated effective PFOS and PFOA removal as reported by the Water Research Foundation (WRF)<ref name="Dickenson2016"/>. Analysis of permeate at both sites referenced by the WRF confirmed that short and long chain PFAS concentrations in the treated water were reduced to levels below test method reporting limits.
 
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Full-scale studies using larger effective pore size NF membranes for PFAS removal are limited in number but are promising since NF systems are somewhat less costly than RO and may be nearly as effective in removing PFAS.  Recent laboratory or pilot studies have shown good performance of NF membranes<ref name="Steinle-Darling2008">Steinle-Darling, E., and Reinhard, M., 2008. Nanofiltration for Trace Organic Contaminant Removal: Structure, Solution, and Membrane Fouling Effects on the Rejection of Perfluorochemicals. Environmental Science and Technology, 42(14), pp. 5292-5297.  [https://doi.org/10.1021/es703207s DOI: 10.1021/es703207s]&nbsp;&nbsp; Free download from: [https://d1wqtxts1xzle7.cloudfront.net/48926882/es703207s20160918-21142-1xmqco5.pdf?1474189169=&response-content-disposition=inline%3B+filename%3DNanofiltration_for_Trace_Organic_Contami.pdf&Expires=1613000850&Signature=N-ZvvjOJX3TSOQzg7od3Q0LulNSZOqqjfummVEUfmiYlC3VasS4FuBHOgY52Xy~7FrKbOLhx0xx8QHdUsR~fbRTMQNXhiqbEslnU2gda2EcZHMMJj0mf-01wIA3jFIywA7IIabmTd3uMUGsIfT1D0PrGY00RmprYIQBoG3Dg~KjoizdfxYfvEgdZw2C~7D47pPiwMSnavZiGuvO0~dbRF8nawL7Prg91xt5BFTNUQQiIrIlMWc4PhVjzE5Su2CUZqnNlYdAW5Ck7B9lKmmVMPiOgz07vFnyp7m-q4UK3woa~aBFW9Wp~hjqN6vfohn8Hocv5oMpZNamhu8vBbPilKw__&Key-Pair-Id=APKAJLOHF5GGSLRBV4ZA Academia].</ref><ref name="Appleman2013">Appleman, T.D., Dickenson, E.R., Bellona, C., and Higgins, C.P., 2013. Nanofiltration and granular activated carbon treatment of perfluoroalkyl acids. Journal of Hazardous Materials, 260, 740-746.  [https://doi.org/10.1016/j.jhazmat.2013.06.033 DOI: 10.1016/j.jhazmat.2013.06.033]</ref><ref name="Soriano2017">Soriano, Á., Gorri, D., and Urtiaga, A., 2017. Efficient treatment of perfluorohexanoic acid by nanofiltration followed by electrochemical degradation of the NF concentrate. Water Research, 112, 147-156.  [https://doi.org/10.1016/j.watres.2017.01.043 DOI: 10.1016/j.watres.2017.01.043]&nbsp;&nbsp; [[Media: Soriano2017.pdf | Author’s Manuscript.]]</ref><ref name="Zeng2017">Zeng, C., Tanaka, S., Suzuki, Y., Yukioka, S., and Fujii, S., 2017. Rejection of Trace Level Perfluorohexanoic Acid (PFHxA) in Pure Water by Loose Nanofiltration Membrane. Journal of Water and Environment Technology, 15(3), pp. 120-127.  [https://doi.org/10.2965/jwet.16-072 DOI: 10.2965/jwet.16-072]&nbsp;&nbsp; Free download from: [https://www.jstage.jst.go.jp/article/jwet/15/3/15_16-072/_pdf J-STAGE]</ref><ref name="Wang2018">Wang, J., Wang, L., Xu, C., Zhi, R., Miao, R., Liang, T., Yue, X., Lv, Y. and Liu, T., 2018. Perfluorooctane sulfonate and perfluorobutane sulfonate removal from water by nanofiltration membrane: The roles of solute concentration, ionic strength, and macromolecular organic foulants. Chemical Engineering Journal, 332, pp. 787-797.  [https://doi.org/10.1016/j.cej.2017.09.061 DOI: 10.1016/j.cej.2017.09.061]</ref>.
 
  
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Although membrane RO and NF processes are generally capable of providing uniform removal rates relative to short and long chain PFAS compounds (see [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | PFAS]] for nomenclature), other aspects of these treatment technologies are more challenging:
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Internationally, Canada and the European Union have also promulgated drinking water standards for select PFAS. However, large discrepancies exist among the various regulatory organizations, largely due to the different effect endpoints and exposure doses being used to calculate risk-based levels. The PFAS guidance from the Interstate Technology and Regulatory Council (ITRC) in the US includes a regularly updated compilation of screening values for PFAS and is available on their PFAS website<ref name="ITRC2023">Interstate Technology and Regulatory Council (ITRC) 2023. PFAS Technical and Regulatory Guidance Document. [https://pfas-1.itrcweb.org/ ITRC PFAS Website]</ref>: https://pfas-1.itrcweb.org.
  
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* Membranes must be flushed and cleaned periodically, such that overall water recovery rates (process water volumes consumed, wasted, and lost vs. treated water volumes produced) are much lower than those for GAC and IX processes. Membrane fouling can be slowed or avoided depending on operating conditions, membrane modifications, and feed modifications<ref name="LeRoux2005">Le Roux, I., Krieg, H.M., Yeates, C.A. and Breytenbach, J.C., 2005. Use of chitosan as an antifouling agent in a membrane bioreactor. Journal of Membrane Science, 248(1-2), pp. 127-136. [https://doi.org/10.1016/j.memsci.2004.10.005 DOI: 10.1016/j.memsci.2004.10.005]</ref>. Typically, 70-90% of the water supplied into a membrane RO process is recoverable as treated water. The remaining 10-30% is reject containing approximately 4 to 8 times the initial PFAS concentration (depending on recovery rate).
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===Ecological Screening Levels===
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Most peer-reviewed literature and regulatory-based environmental quality benchmarks have been developed using data for PFOS and PFOA; however, other select PFAAs have been evaluated for potential effects to aquatic receptors<ref name="ITRC2023"/><ref name="ZodrowEtAl2021a"/><ref name="ConderEtAl2020"/>. USEPA has developed water quality criteria for aquatic life<ref name="USEPA2022"> United States Environmental Protection Agency (USEPA), 2022. Fact Sheet: Draft 2022 Aquatic Life Ambient Water Quality Criteria for Perfluorooctanoic acid (PFOA) and Perfluorooctane Sulfonic Acid (PFOS)). Office of Water, EPA 842-D-22-005. [[Media: USEPA2022.pdf | Fact Sheet]]</ref><ref name="USEPA2024c">United States Environmental Protection Agency (USEPA), 2024. Final Freshwater Aquatic Life Ambient Water Quality Criteria and Acute Saltwater Aquatic Life Benchmark for Perfluorooctanoic Acid (PFOA). Office of Water, EPA-842-R-24-002. [[Media: USEPA2024c.pdf | Report.pdf]]</ref><ref name="USEPA2024d">United States Environmental Protection Agency (USEPA), 2024. Final Freshwater Aquatic Life Ambient Water Quality Criteria and Acute Saltwater Aquatic Life Benchmark for Perfluorooctane Sulfonate (PFOS). Office of Water, EPA-842-R-24-003. [[Media: USEPA2024d.pdf | Report.pdf]]</ref> for PFOA and PFOS. Following extensive reviews of the peer-reviewed literature, Zodrow ''et al.''<ref name="ZodrowEtAl2021a"/> used the USEPA Great Lakes Initiative methodology<ref>United States Environmental Protection Agency (USEPA), 2012. Water Quality Guidance for the Great Lakes System. Part 132. [https://www.govinfo.gov/app/details/CFR-2013-title40-vol23/CFR-2013-title40-vol23-part132 Government Website]&nbsp; [[Media: CFR-2013-title40-vol23-part132.pdf | Part132.pdf]]</ref> to calculate acute and chronic screening levels for aquatic life for 23 PFAS. The Argonne National Laboratory has also developed Ecological Screening Levels for multiple PFAS<ref name="GrippoEtAl2024">Grippo, M., Hayse, J., Hlohowskyj, I., Picel, K., 2024. Derivation of PFAS Ecological Screening Values - Update. Argonne National Laboratory Environmental Science Division. [[Media: GrippoEtAl2024.pdf | Report.pdf]]</ref>. In contrast to surface water aquatic life benchmarks, sediment benchmark values are limited. For terrestrial systems, screening levels for direct exposure of soil plants and invertebrates to PFAS in soils have been developed for multiple AFFF-related PFAS<ref name="ConderEtAl2020"/><ref name="ZodrowEtAl2021a"/>, and the Canadian Council of Ministers of Environment developed several draft thresholds protective of direct toxicity of PFOS in soil<ref>Canadian Council of Ministers of the Environment (CCME), 2021. Canadian Soil and Groundwater Quality Guidelines for the Protection of Environmental and Human Health, Perfluorooctane Sulfonate (PFOS). [[Media: CCME2018.pdf | Open Access Government Document]]</ref>.  
  
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* These cleaning and flushing processes create a continuous liquid waste stream, which periodically includes harsh membrane cleaning chemicals as well as a continuous flow of concentrated membrane reject chemicals (i.e., PFAS) that must be properly managed and disposed of. Management often includes further treatment to remove PFAS from the liquid waste.
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Wildlife screening levels for abiotic media are back-calculated from food web models developed for representative receptors. Both Zodrow ''et al.''<ref name="ZodrowEtAl2021a"/> and Grippo ''et al.''<ref name="GrippoEtAl2024"/> include the development of risk-based screening levels for wildlife. The Michigan Department of Community Health<ref>Dykema, L.D., 2015. Michigan Department of Community Health Final Report, USEPA Great Lakes Restoration Initiative (GLRI) Project, Measuring Perfluorinated Compounds in Michigan Surface Waters and Fish. Grant GL-00E01122. [https://www.michigan.gov/documents/mdch/MDCH_GL-00E01122-0_Final_Report_493494_7.pdf Free Download]&nbsp; [[Media: MDCH_Geart_Lakes_PFAS.pdf | Report.pdf]]</ref> derived a provisional PFOS surface water value for avian and mammalian wildlife. In California, the San Francisco Bay Regional Water Quality Control Board developed terrestrial habitat soil ecological screening levels based on values developed in Zodrow ''et al.''<ref name="ZodrowEtAl2021a"/>. For PFOS only, a dietary screening level (i.e. applicable to the concentration of PFAS measured in dietary items) has been developed for mammals at 4.6 micrograms per kilogram (μg/kg) wet weight (ww), and for avians at 8.2 μg/kg ww<ref>Environment and Climate Change Canada, 2018. Federal Environmental Quality Guidelines, Perfluorooctane Sulfonate (PFOS). [[Media: ECCC2018.pdf | Repoprt.pdf]]</ref>.
  
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* RO and NF systems are inherently more expensive and complicated systems to implement, operate, and maintain compared to adsorption processes. Treatment system operator certification and process monitoring requirements are correspondingly markedly higher for RO and NF than they are for GAC and IX.  
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==Approaches for Evaluating Exposures and Effects in AFFF Site Environmental Risk Assessment: Human Health==
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Exposure pathways and effects for select PFAS are well understood, such that standard human health risk assessment approaches can be used to quantify risks for populations relevant to a site. Human health exposures via drinking water have been the focus in risk assessments and investigations at PFAS sites<ref>Post, G.B., Cohn, P.D., Cooper, K.R., 2012. Perfluorooctanoic acid (PFOA), an emerging drinking water contaminant: A critical review of recent literature. Environmental Research, 116, pp. 93-117. [https://doi.org/10.1016/j.envres.2012.03.007 doi: 10.1016/j.envres.2012.03.007]</ref><ref>Guelfo, J.L., Marlow, T., Klein, D.M., Savitz, D.A., Frickel, S., Crimi, M., Suuberg, E.M., 2018. Evaluation and Management Strategies for Per- and Polyfluoroalkyl Substances (PFASs) in Drinking Water Aquifers: Perspectives from Impacted U.S. Northeast Communities. Environmental Health Perspectives,126(6), 13 pages. [https://doi.org/10.1289/EHP2727 doi: 10.1289/EHP2727]&nbsp; [[Media: GuelfoEtAl2018.pdf | Open Access Article]]</ref>. Risk assessment approaches for PFAS in drinking water follow typical, well-established drinking water risk assessment approaches for chemicals as detailed in regulatory guidance documents for various jurisdictions.  
  
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* Water feed pressures required to drive flow through membrane RO and NF processes are considerably higher than those involved with GAC and IX processes. This results in reduced process efficiency and higher pumping and electrical operating costs.
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Incidental exposures to soil and dusts for PFAS can occur during a variety of soil disturbance activities, such as gardening and digging, hand-to-mouth activities, and intrusive groundwork by industrial or construction workers. As detailed by the ITRC<ref name="ITRC2023"/>, many US states and USEPA have calculated risk-based screening levels for these soil and drinking water pathways (and many also include dermal exposures to soils) using well-established risk assessment guidance.  
  
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* Membrane systems can also be subject to issues with irreversible membrane fouling, clogging, and scaling or other physical membrane damage and failures. Additional water pretreatment and higher levels of monitoring and maintenance are then required, further adding to the higher costs of such systems.
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Field and laboratory studies have shown that some PFCAs and PFSAs bioaccumulate in fish and other aquatic life at rates that could result in relevant dietary PFAS exposures for consumers of fish and other seafood<ref>Martin, J.W., Mabury, S.A., Solomon, K.R., Muir, D.C., 2003. Dietary accumulation of perfluorinated acids in juvenile rainbow trout (Oncorhynchus mykiss). Environmental Toxicology and Chemistry, 22(1), pp.189-195. [https://doi.org/10.1002/etc.5620220125 doi: 10.1002/etc.5620220125]</ref><ref>Martin, J.W., Mabury, S.A., Solomon, K.R., Muir, D.C., 2003. Bioconcentration and tissue distribution of perfluorinated acids in rainbow trout (Oncorhynchus mykiss). Environmental Toxicology and Chemistry, 22(1), pp.196-204. [https://doi.org/10.1002/etc.5620220126 doi: 10.1002/etc.5620220126]</ref><ref>Chen, F., Gong, Z., Kelly, B.C., 2016. Bioavailability and bioconcentration potential of perfluoroalkyl-phosphinic and -phosphonic acids in zebrafish (Danio rerio): Comparison to perfluorocarboxylates and perfluorosulfonates. Science of The Total Environment, 568, pp. 33-41. [https://doi.org/10.1016/j.scitotenv.2016.05.215 doi: 10.1016/j.scitotenv.2016.05.215]</ref><ref>Fang, S., Zhang, Y., Zhao, S., Qiang, L., Chen, M., Zhu, L., 2016. Bioaccumulation of per fluoroalkyl acids including the isomers of perfluorooctane sulfonate in carp (Cyprinus carpio) in a sediment/water microcosm. Environmental Toxicology and Chemistry, 35(12), pp. 3005-3013. [https://doi.org/10.1002/etc.3483 doi: 10.1002/etc.3483]</ref><ref>Bertin, D., Ferrari, B.J.D. Labadie, P., Sapin, A., Garric, J., Budzinski, H., Houde, M., Babut, M., 2014. Bioaccumulation of perfluoroalkyl compounds in midge (Chironomus riparius) larvae exposed to sediment. Environmental Pollution, 189, pp. 27-34. [https://doi.org/10.1016/j.envpol.2014.02.018  doi: 10.1016/j.envpol.2014.02.018]</ref><ref>Bertin, D., Labadie, P., Ferrari, B.J.D., Sapin, A., Garric, J., Geffard, O., Budzinski, H., Babut. M., 2016. Potential exposure routes and accumulation kinetics for poly- and perfluorinated alkyl compounds for a freshwater amphipod: Gammarus spp. (Crustacea). Chemosphere, 155, pp. 380-387. [https://doi.org/10.1016/j.chemosphere.2016.04.006 doi: 10.1016/j.chemosphere.2016.04.006]</ref><ref>Dai, Z., Xia, X., Guo, J., Jiang, X., 2013. Bioaccumulation and uptake routes of perfluoroalkyl acids in Daphnia magna. Chemosphere, 90(5), pp.1589-1596. [https://doi.org/10.1016/j.chemosphere.2012.08.026 doi: 10.1016/j.chemosphere.2012.08.026]</ref><ref>Prosser, R.S., Mahon, K., Sibley, P.K., Poirier, D., Watson-Leung, T. 2016. Bioaccumulation of perfluorinated carboxylates and sulfonates and polychlorinated biphenyls in laboratory-cultured Hexagenia spp., Lumbriculus variegatus and Pimephales promelas from field-collected sediments. Science of The Total Environment, 543(A), pp. 715-726. [https://doi.org/10.1016/j.scitotenv.2015.11.062 doi: 10.1016/j.scitotenv.2015.11.062]</ref><ref>Rich, C.D., Blaine, A.C., Hundal, L., Higgins, C., 2015. Bioaccumulation of Perfluoroalkyl Acids by Earthworms (Eisenia fetida) Exposed to Contaminated Soils. Environmental Science and Technology, 49(2) pp. 881-888. [https://doi.org/10.1021/es504152d doi: 10.1021/es504152d]</ref><ref>Muller, C.E., De Silva, A.O., Small, J., Williamson, M., Wang, X., Morris, A., Katz, S., Gamberg, M., Muir, D.C.G., 2011. Biomagnification of Perfluorinated Compounds in a Remote Terrestrial Food Chain: Lichen–Caribou–Wolf. Environmental Science and Technology, 45(20), pp. 8665-8673. [https://doi.org/10.1021/es201353v doi: 10.1021/es201353v]</ref>. In addition to fish, terrestrial wildlife can accumulate contaminants from impacted sites, resulting in potential exposures to consumers of wild game<ref name="ConderEtAl2021"/>. Additionally, exposures can occur though consumption of homegrown produce or agricultural products that originate from areas irrigated with PFAS-impacted groundwater, or that are amended with biosolids that contain PFAS, or that contain soils that were directly affected by PFAS releases<ref>Brown, J.B, Conder, J.M., Arblaster, J.A., Higgins, C.P.,  2020. Assessing Human Health Risks from Per- and Polyfluoroalkyl Substance (PFAS)-Impacted Vegetable Consumption: A Tiered Modeling Approach. Environmental Science and Technology, 54(23), pp. 15202-15214. [https://doi.org/10.1021/acs.est.0c03411 doi: 10.1021/acs.est.0c03411]&nbsp; [[Media: BrownEtAl2020.pdf | Open Access Article]]</ref>. Multiple studies have found PFAS can be taken up by plants from soil porewater<ref>Blaine, A.C., Rich, C.D., Hundal, L.S., Lau, C., Mills, M.A., Harris, K.M., Higgins, C.P., 2013. Uptake of Perfluoroalkyl Acids into Edible Crops via Land Applied Biosolids: Field and Greenhouse Studies. Environmental Science and Technology, 47(24), pp. 14062-14069. [https://doi.org/10.1021/es403094q doi: 10.1021/es403094q]&nbsp; [https://www.epa.gov/sites/production/files/2019-11/documents/508_pfascropuptake.pdf Free Download from epa.gov]</ref><ref>Blaine, A.C., Rich, C.D., Sedlacko, E.M., Hyland, K.C., Stushnoff, C., Dickenson, E.R.V., Higgins, C.P., 2014. Perfluoroalkyl Acid Uptake in Lettuce (Lactuca sativa) and Strawberry (Fragaria ananassa) Irrigated with Reclaimed Water. Environmental Science and Technology, 48(24), pp. 14361-14368. [https://doi.org/10.1021/es504150h doi: 10.1021/es504150h]</ref><ref>Ghisi, R., Vamerali, T., Manzetti, S., 2019. Accumulation of perfluorinated alkyl substances (PFAS) in agricultural plants: A review. Environmental Research, 169, pp. 326-341. [https://doi.org/10.1016/j.envres.2018.10.023 doi: 10.1016/j.envres.2018.10.023]</ref>, and livestock can accumulate PFAS from drinking water and/or feed<ref>van Asselt, E.D., Kowalczyk, J., van Eijkeren, J.C.H., Zeilmaker, M.J., Ehlers, S., Furst, P., Lahrssen-Wiederhold, M., van der Fels-Klerx, H.J., 2013. Transfer of perfluorooctane sulfonic acid (PFOS) from contaminated feed to dairy milk. Food Chemistry, 141(2), pp.1489-1495. [https://doi.org/10.1016/j.foodchem.2013.04.035 doi: 10.1016/j.foodchem.2013.04.035]</ref>. Thus, when PFAS are present in surface water bodies where fishing or shellfish harvesting occurs or terrestrial areas where produce is grown or game is hunted, the bioaccumulation of PFAS into dietary items can be an important pathway for human exposure.  
  
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===Activated Carbon Adsorption===
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PFAAs such as PFOA and PFOS are not expected to volatilize from PFAS-impacted environmental media<ref name="USEPA2016a"/><ref name="USEPA2016b"/> such as soil and groundwater, which are the primary focus of most site-specific risk assessments. In contrast to non-volatile PFAAs, fluorotelomer alcohols (FTOHs) are among the more widely studied of the volatile PFAS. FTOHs are transient in the atmosphere with a lifetime of 20 days<ref>Ellis, D.A., Martin, J.W., De Silva, A.O., Mabury, S.A., Hurley, M.D., Sulbaek Andersen, M.P., Wallington, T.J., 2004. Degradation of Fluorotelomer Alcohols:  A Likely Atmospheric Source of Perfluorinated Carboxylic Acids. Environmental Science and Technology, 38(12), pp. 3316-3321. [https://doi.org/10.1021/es049860w doi: 10.1021/es049860w]</ref>. At most AFFF sites under evaluation, AFFF releases have occurred many years before such that FTOH may no longer be present. As such, the current assumption is that volatile PFAS, such as FTOHs historically released at the site, will have transformed to stable, low-volatility PFAS, such as PFAAs in soil or groundwater, or will they have diffused to the outdoor atmosphere. There is no evidence that FTOHs or other volatile PFAS are persistent in groundwater or soils such that they present an indoor vapor intrusion pathway risk concern as observed for chlorinated solvents. Ongoing research continues for the vapor pathway<ref name="ITRC2023"/>.
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[[File: GAChouse.JPG | thumb| 500px | Figure 2.  Typical private water supply well GAC installation for removal PFAS. Pressure gages and sample ports installed before the first (or lead) vessel, at the midpoint, and after the second (or lag) vessel allow monitoring for pressure drop due to fouling and for contaminant breakthrough.]] 
 
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Activated carbon is a form of carbon processed to have small pores that increase the surface area available for adsorption of constituents from water. Activated carbon is derived from many source materials, including coconut shells, wood, lignite, and bituminous coal. Different types of activated carbon base materials have varied adsorption characteristics such that some may be better suited to removing certain contaminant compounds than others. Results from laboratory testing, pilot evaluations, and full-scale system operations suggest that bituminous coal-based GAC is generally the best performing carbon for PFAS removal<ref name="McNamara2018">McNamara, J.D., Franco, R., Mimna, R., and Zappa, L., 2018. Comparison of Activated Carbons for Removal of Perfluorinated Compounds from Drinking Water. Journal‐American Water Works Association, 110(1), pp. E2-E14. [https://doi.org/10.5942/jawwa.2018.110.0003 DOI: 10.5942/jawwa.2018.110.0003]</ref><ref name="Westreich2018">Westreich, P., Mimna, R., Brewer, J., and Forrester, F., 2018. The removal of short‐chain and long‐chain perfluoroalkyl acids and sulfonates via granular activated carbons: A comparative column study. Remediation Journal, 29(1), pp. 19-26. [https://doi.org/10.1002/rem.21579 DOI: 10.1002/rem.21579]</ref>.
 
  
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The removal efficiency of individual PFAS compounds using GAC is a function of both the PFAS functional group (carboxylic acid versus sulfonic acid) and also the perfluoro-carbon chain length<ref name="McCleaf2017">McCleaf, P., Englund, S., Östlund, A., Lindegren, K., Wiberg, K., and Ahrens, L., 2017. Removal efficiency of multiple poly-and perfluoroalkyl substances (PFASs) in drinking water using granular activated carbon (GAC) and anion exchange (AE) column tests. Water Research, 120, pp. 77-87. [https://doi.org/10.1016/j.watres.2017.04.057 DOI: 10.1016/j.watres.2017.04.057]</ref><ref name="Eschauzier2012">Eschauzier, C., Beerendonk, E., Scholte-Veenendaal, P., and De Voogt, P., 2012. Impact of Treatment Processes on the Removal of Perfluoroalkyl Acids from the Drinking Water Production Chain. Environmental Science and Technology, 46(3), pp. 1708-1715. [https://doi.org/10.1021/es201662b DOI: 10.1021/es201662b]</ref>(see [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | PFAS]] for nomenclature):
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General and site-specific human health exposure pathways and risk assessment methods as outlined by USEPA<ref>United States Environmental Protection Agency (USEPA), 1989. Risk Assessment Guidance for Superfund: Volume I, Human Health Evaluation Manual (Part A). Office of Solid Waste and Emergency Response, EPA/540/1-89/002. [https://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=10001FQY.txt Free Download]&nbsp; [[Media: USEPA1989.pdf | Report.pdf]]</ref><ref name="USEPA1997">United States Environmental Protection Agency (USEPA), 1997. Ecological Risk Assessment Guidance for Superfund: Process for Designing and Conducting Ecological Risk Assessments, Interim Final. Office of Solid Waste and Emergency Response, EPA 540-R-97-006. [http://semspub.epa.gov/src/document/HQ/157941 Free Download]&nbsp; [[Media: EPA540-R-97-006.pdf | Report.pdf]]</ref> can be applied to PFAS risk assessments for which human health toxicity values have been developed. Additionally, for risk assessments with dietary exposures of PFAS, standard risk assessment food web modeling can be used to develop initial estimates of dietary concentrations which can be confirmed with site-specific tissue sampling programs.
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* perfluoro-sulfonate acids (PFSAs) are more efficiently removed than perfluoro-carboxylic acids (PFCAs) of the same chain length
 
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* long chain compounds of the same functional group are removed better than the shorter chains
 
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Activated carbon may be applied in drinking water systems as GAC or PAC<ref name="Dudley">Dudley, L.A., Arevalo, E.C., and Knappe, D.R., 2015. Removal of Perfluoroalkyl Substances by PAC Adsorption and Anion Exchange. Water Research Foundation Project #4344.  Free  download of Executive Summary from: [https://www.waterrf.org/system/files/resource/2019-04/4344_ProjectSummary.pdf Water Research Foundation (Public Plus account)]</ref><ref name="Qian2017">Qian, J., Shen, M., Wang, P., Wang, C., Li, K., Liu, J., Lu, B. and Tian, X., 2017. Perfluorooctane sulfonate adsorption on powder activated carbon: Effect of phosphate (P) competition, pH, and temperature. Chemosphere, 182, pp. 215-222.  [https://doi.org/10.1016/j.chemosphere.2017.05.033 DOI: 10.1016/j.chemosphere.2017.05.033]</ref>. GAC has larger granules and is reusable, while PAC has much smaller granules and is not typically reused.  To-date, PAC is most often used as a temporary treatment as costs associated with disposal and replacement of the used PAC may preclude using PAC for long-term treatment. A typical GAC installation for a private drinking water well is shown in Figure 2.
 
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Contrary to PAC, GAC used to treat PFAS can be reactivated by the manufacturer, driving the PFAS from the GAC and into off-gas. The extracted gas is then treated with thermal oxidation (temperatures often 1200&deg;C to 1400&deg;C).  The reactivated GAC is then brought back to the site and reused.  Thus, GAC can ultimately be a destructive treatment technology.  
 
  
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[[File: IXcycle.png | thumb | 400px | left | Figure 3.  Operational cycle of a packed bed reactor with anion exchange resin beads]]
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==Approaches for Evaluating Exposures and Effects in AFFF Site Environmental Risk Assessment: Ecological==
 +
Information available currently on exposures and effects of PFAS in ecological receptors indicate that the PFAS ecological risk issues at most sites are primarily associated with risks to vertebrate wildlife.  Avian and mammalian wildlife are relatively sensitive to PFAS, and dietary intake via bioaccumulation in terrestrial and aquatic food webs can result in exposures that are dominated by the more accumulative PFAS<ref name="LarsonEtAl2018">Larson, E.S., Conder, J.M., Arblaster, J.A., 2018. Modeling avian exposures to perfluoroalkyl substances in aquatic habitats impacted by historical aqueous film forming foam releases. Chemosphere, 201, pp. 335-341. [https://doi.org/10.1016/j.chemosphere.2018.03.004 doi: 10.1016/j.chemosphere.2018.03.004]</ref><ref name="ConderEtAl2020"/><ref name="ZodrowEtAl2021a"/>. Direct toxicity to aquatic life (e.g., fish, pelagic life, benthic invertebrates, and aquatic plants) can occur from exposure to sediment and surface water at effected sites.  For larger areas, surface water concentrations associated with adverse effects to aquatic life are generally higher than those that could result in adverse effects to aquatic-dependent wildlife. Soil invertebrates and plants are generally less sensitive, with risk-based concentrations in soil being much higher than those associated with potential effects to terrestrial wildlife<ref name="ZodrowEtAl2021a"/>.
  
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===Anion Exchange===
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Aquatic life are exposed to PFAS through direct exposure in surface water and sediment. Ecological risk assessment approaches for PFAS for aquatic life follow standard risk assessment approaches. The evaluation of potential risks for aquatic life with direct exposure to PFAS in environmental media relies on comparing concentrations in external exposure media to protective, media-specific benchmarks, including the aquatic life risk-based screening levels discussed above<ref name="ZodrowEtAl2021a"/><ref name="USEPA2024a">United States Environmental Protection Agency (USEPA), 2024. National Recommended Water Quality Criteria - Aquatic Life Criteria Table. [https://www.epa.gov/wqc/national-recommended-water-quality-criteria-aquatic-life-criteria-table USEPA Website]</ref>.
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Anion exchange has also been demonstrated for the adsorption of PFAS, and published results note higher sorption per pound than GAC<ref name="McCleaf2017"/><ref name=" Senevirathna2010">Senevirathna, S.T.M.L.D., Tanaka, S., Fujii, S., Kunacheva, C., Harada, H., Shivakoti, B.R., and Okamoto, R., 2010. A comparative study of adsorption of perfluorooctane sulfonate (PFOS) onto granular activated carbon, ion-exchange polymers and non-ion-exchange polymers. Chemosphere, 80(6), pp. 647-651.  [https://doi.org/10.1016/j.chemosphere.2010.04.053 DOI: 10.1016/j.chemosphere.2010.04.053]&nbsp;&nbsp; Free download from: [https://www.researchgate.net/profile/Chinagarn_Kunacheva/publication/44672056_A_comparative_study_of_adsorption_of_perfluorooctane_sulfonate_PFOS_onto_granular_activated_carbon_ion-exchange_polymers_and_non-ion-exchange_polymers/links/5a3380510f7e9b2a288a2b21/A-comparative-study-of-adsorption-of-perfluorooctane-sulfonate-PFOS-onto-granular-activated-carbon-ion-exchange-polymers-and-non-ion-exchange-polymers.pdf ResearchGate]</ref><ref name="Woodard2017">Woodard, S., Berry, J., and Newman, B., 2017. Ion exchange resin for PFAS removal and pilot test comparison to GAC. Remediation Journal, 27(3), pp. 19-27.  [https://doi.org/10.1002/rem.21515 DOI: 10.1002/rem.21515]</ref>. The higher capacity is believed to be due to combined hydrophobic and ion exchange adsorption mechanisms, whereas GAC mainly relies on hydrophobic attraction. Anion exchange resins can be highly selective, or they can also remove other contaminants based on design requirements and water chemistry. Resins have greater affinity for PFAS subgroup PFSA than for PFCA, and affinity increases with carbon chain length.
 
−
[[Wikipedia: Ion-exchange resin | Anion exchange resins]] are a viable alternative to GAC for ex situ treatment of PFAS anions, and several venders sell resins capable of removing PFAS. Resins available for treating PFAS include regenerable resins that can be used multiple times (Figure 3) and single-use resins that must be disposed or destroyed after use<ref name=" Senevirathna2010"/>. Regenerable resins generate a solvent and brine solution, which is distilled to recover the solvent prior to the brine being adsorbed onto a small quantity of GAC or resin for ultimate disposal. This use of one treatment technology (GAC, IX) to support another (RO) is sometimes referred to as a “treatment train” approach. Single-use resins can be more fully exhausted than regenerable resins can and may be a more cost-effective solution for low concentration PFAS contamination, while regenerable resins may be more cost effective for high concentration contamination.
 
  
−
==Developing PFAS Treatment Technologies==
+
When an area at the point of PFAS release is an industrial setting which does not feature favorable habitats for terrestrial and aquatic-dependent wildlife, the transport mechanisms may allow PFAS to travel offsite. If offsite or downgradient areas contain ecological habitat, then PFAS transported to these areas are expected to pose the highest risk potential to wildlife, particularly those areas that feature aquatic habitat<ref>Ahrens, L., Bundschuh, M., 2014. Fate and effects of poly- and perfluoroalkyl substances in the aquatic environment: A review. Environmental Toxicology and Chemistry, 33(9), pp. 1921-1929. [https://doi.org/10.1002/etc.2663 doi: 10.1002/etc.2663]&nbsp; [[Media: AhrensBundschuh2014.pdf | Open Access Article]]</ref><ref name="LarsonEtAl2018"/>.
−
Numerous separation and destructive technologies are in the developmental stages of bench-scale testing or limited field-scale demonstrations. Some of these are listed in Table 1:  
 
  
−
{| class="wikitable" style="float:right; margin-left:10px;"
+
Wildlife receptors, specifically birds and mammals, are typically exposed to PFAS through uptake from dietary sources such as plants and invertebrates, along with direct soil ingestion during foraging activities. Dietary intake modeling typical for ecological risk assessments is the recommended approach for an evaluation of potential risks to wildlife species where PFAS exposure occurs primarily via dietary uptake from bioaccumulation pathways. Dietary intake modeling uses relevant exposure factors for each receptor group (terrestrial birds, terrestrial mammals, aquatic-dependent birds, and aquatic mammals) to determine a total daily intake (TDI) of PFAS via all potential exposure pathways. This approach requires determination of concentrations of PFAS in dietary items, which can be obtained by measuring PFAS in biota at sites or by using food web models to predict concentrations in biota using measured concentrations of PFAS in soil, sediment, or surface water. Food web models use bioaccumulation metrics such as bioaccumulation factors (BAFs) and biomagnification factors (BMFs) with measurements of PFAS in abiotic media to estimate concentrations in dietary items, including plants and benthic or pelagic invertebrates, to model wildlife exposure and calculate TDI. Once site-specific TDI values are calculated, they are compared to known TRVs identified from toxicity data with exposure doses associated with a lack of adverse effects (termed no observed adverse effect level [NOAEL]) or low adverse effects (termed lowest observed adverse effect level [LOAEL]), per standard risk assessment practice<ref name="USEPA1997"/>.
−
|+ Table 1. Developmental Technologies
 
−
|-
 
−
! Stage
 
−
! Separation/Transfer
 
−
! Destructive*
 
−
|-
 
−
| Developing
 
−
|
 
−
* Biochar (20, 21, 22)
 
−
* Modified Zeolites (23, 24)
 
−
* Specialty adsorbents
 
−
|
 
−
* Electro-oxidation (32, 33, 34)
 
−
* Heat activated persulfate (35)
 
−
* Alkaline perozone (36)
 
−
* Sonolysis (37, 38, 39, 40)
 
−
* Super Critical Water Oxidation
 
−
|-
 
−
| Maturing and</br>Demonstrated
 
−
|
 
−
* Chemical coagulation (28)
 
−
* Electrocoagulation (29)
 
−
* Foam fractionation (30, 31)
 
−
|
 
−
* Low temperature plasma (41, 42)
 
−
|-
 
−
| colspan="3" style="background:white;" | * There are several other destructive technologies such as alternative oxidants, and activation</br>methods of oxidants, but for the purpose of this article, the main categories are presented here.
 
−
|}
 
  
−
==Conclusions==
+
Recently, Conder ''et al.''<ref name="ConderEtAl2020"/>, Gobas ''et al.''<ref name="GobasEtAl2020"/>, and Zodrow ''et al.''<ref name="ZodrowEtAl2021a"/> compiled bioaccumulation modeling parameters and approaches for terrestrial and aquatic food web modeling of a variety of commonly detected PFAS at AFFF sites. There are also several sources of TRVs which can be relied upon for estimating TDI values<ref name="ConderEtAl2020"/><ref name="GobasEtAl2020"/><ref name="ZodrowEtAl2021a"/><ref>Newsted, J.L., Jones, P.D., Coady, K., Giesy, J.P., 2005. Avian Toxicity Reference Values for Perfluorooctane Sulfonate. Environmental Science and Technology, 39(23), pp. 9357-9362. [https://doi.org/10.1021/es050989v doi: 10.1021/es050989v]</ref><ref name="Suski2020"/>. In general, the highest risk for PFAS is expected for smaller insectivore and omnivore receptors (e.g., shrews and other small rodents, small nonmigratory birds), which tend to be lower in trophic level and spend more time foraging in small areas similar to or smaller in size than the impacted area. Compared to smaller, lower-trophic level organisms, larger mammalian and avian carnivores are expected to have lower exposures from site-specific PFAS sources because they forage over larger areas that may include areas that are not impacted, as compared to small organisms with small home ranges<ref name="LarsonEtAl2018"/><ref name="ConderEtAl2020"/><ref name="GobasEtAl2020"/><ref name="Suski2020"/><ref name="ZodrowEtAl2021a"/>.
−
The well established processes for removing PFAS from water all produce residuals that require management, and it is likely that newer processes under development will also produce some residuals. Often, it is the residuals that limit the usefulness of the process.  For instance, RO and NF may currently provide the most complete treatment of water, but the production of a relatively high volume of PFAS-containing liquid reject (the portion of the liquid that retains the contaminants and is “rejected” from the process) limits their application. Often, a second treatment technology such as an adsorbent is required to support the main technology by concentrating or treating the residuals.
 
−
As more testing and operational data on adsorbents are generated, it is becoming evident that no adsorbent technology outperforms the others in all cases. Whether GAC, ion exchange or another technology is the most technically efficient and cost effective long term option for a given site depends on influent water geochemistry and contaminant concentrations, treatment standards, co-contaminants, duration of treatment, and required flow rates. New generation adsorbents are rapidly being introduced into the market at “evaluation scale” which may provide advantages over commercially available adsorbents.
 
−
Several newer technologies are being evaluated in the lab and in the field which include electro-oxidation, heat-activated persulfate, sonolysis, electrocoagulation, low temperature plasma, super critical water oxidation, and foam fractionation. These and other potential treatments for PFAS are still largely in the developmental stage. Several technologies show promise for improved management of PFAS sites. However, it is unlikely that a single technology will be adequate for full remediation at many sites. A multi-technology treatment train approach may be necessary for effective treatment of this complicated group of compounds.
 
  
−
<br clear="left" />
+
Available information regarding PFAS exposure pathways and effects in aquatic life, terrestrial invertebrates and plants, as well as aquatic and terrestrial wildlife allow ecological risk assessment methods to be applied as outlined by USEPA<ref name="USEPA1997"/> to site-specific PFAS risk assessments. Additionally, food web modeling can be used in site-specific PFAS risk assessment to develop initial estimates of dietary concentrations for aquatic and terrestrial wildlife, which can be confirmed with tissue sampling programs at a site.
 +
 
 +
==PFAS Risk Assessment Data Gaps==
 +
There are a number of data gaps currently associated with PFAS risk assessment including the following:
 +
*'''Unmeasured PFAS:''' There are a number of additional PFAS that we know little about and many PFAS that we are unable to quantify in the environment. The approach to dealing with the lack of information on the overwhelming number of PFAS is being debated; in the meantime, however, PFAS beyond PFOS and PFOA are being studied more, and this information will result in improved characterization of risks for other PFAS. 
 +
 
 +
*'''Mixtures:''' Another major challenge in effects assessment for PFAS, for both human health risk assessments and environmental risk assessments, is understanding the potential importance of mixtures of PFAS. Considering the limited human health and ecological toxicity data available for just a few PFAS, the understanding of the relative toxicity, additivity, or synergistic effects of PFAS in mixtures is just beginning.
 +
 
 +
*'''Toxicity Data Gaps:''' For environmental risk assessments, some organisms such as reptiles and benthic invertebrates do not have toxicity data available. Benchmark or threshold concentrations of PFAS in environmental media intended to be protective of wildlife and aquatic organisms suffer from significant uncertainty in their derivation due to the limited number of species for which data are available. As species-specific data becomes available for more types of organisms, the accuracy of environmental risk assessments is likely to improve. 
  
 
==References==
 
==References==
−
 
 
<references />
 
<references />
  
 
==See Also==
 
==See Also==
 +
[https://www.atsdr.cdc.gov/pfas/health-studies/index.html Agency for Toxic Substances and Disease Registry (ATSDR) PFAS Health Studies]

Latest revision as of 17:53, 21 October 2025

Remediation of Stormwater Runoff Contaminated by Munition Constituents

Past and ongoing military operations have resulted in contamination of surface soil with munition constituents (MC), which have human and environmental health impacts. These compounds can be transported off site via stormwater runoff during precipitation events. Technologies to “trap and treat” surface runoff before it enters downstream receiving bodies (e.g., streams, rivers, ponds) (see Figure 1), and which are compatible with ongoing range activities are needed. This article describes a passive and sustainable approach for effective management of munition constituents in stormwater runoff.

Related Article(s):


Contributor: Mark E. Fuller

Key Resource(s):

  • SERDP Project ER19-1106: Development of Innovative Passive and Sustainable Treatment Technologies for Energetic Compounds in Surface Runoff on Active Ranges

Background

Surface Runoff Characteristics and Treatment Approaches

Figure 1. Conceptual model of passive trap and treat approach for MC removal from stormwater runoff

During large precipitation events the rate of water deposition exceeds the rate of water infiltration, resulting in surface runoff (also called stormwater runoff). Surface characteristics including soil texture, presence of impermeable surfaces (natural and artificial), slope, and density and type of vegetation all influence the amount of surface runoff from a given land area. The use of passive systems such as retention ponds and biofiltration cells for treatment of surface runoff is well established for urban and roadway runoff. Treatment in those cases is typically achieved by directing runoff into and through a small constructed wetland, often at the outlet of a retention basin, or via filtration by directing runoff through a more highly engineered channel or vault containing the treatment materials. Filtration based technologies have proven to be effective for the removal of metals, organics, and suspended solids[1][2][3][4].

Surface Runoff on Ranges

Figure 2. Conceptual illustration of munition constituent production and transport on military ranges. Mesoscale residues are qualitatively defined as being easily visible to the naked eye (e.g., from around 50 µm to multiple cm in size) and less likely to be transported by moving water. Microscale residues are defined as <50 µm down to below 1 µm, and more likely to be entrained in, and transported by, moving water as particulates. Blue arrows represent possible water flow paths and include both dissolved and solid phase energetics. The red vertical arrow represents the predominant energetics dissolution process in close proximity to the residues due to precipitation.

Surface runoff represents a major potential mechanism through which energetics residues and related materials are transported off site from range soils to groundwater and surface water receptors (Figure 2). This process is particularly important for energetics that are water soluble (e.g., NTO and NQ) or generate soluble daughter products (e.g., DNAN and TNT). While traditional MC such as RDX and HMX have limited aqueous solubility, they also exhibit recalcitrance to degrade under most natural conditions. RDX and perchlorate are frequent groundwater contaminants on military training ranges. While actual field measurements of energetics in surface runoff are limited, laboratory experiments have been performed to predict mobile energetics contamination levels based on soil mass loadings[5][6][7][8][9]. For example, in a previous small study, MC were detected in surface runoff from an active live-fire range[10], and more recent sampling has detected MC in marsh surface water adjacent to the same installation (personal communication). Another recent report from Canada also detected RDX in both surface runoff and surface water at low part per billion levels in a survey of several military demolition sites[11]. However, overall, data regarding the MC contaminant profile of surface runoff from ranges is very limited, and the possible presence of non-energetic constituents (e.g., metals, binders, plasticizers) in runoff has not been examined. Additionally, while energetics-contaminated surface runoff is an important concern, mitigation technologies specifically for surface runoff have not yet been developed and widely deployed in the field. To effectively capture and degrade MC and associated compounds that are present in surface runoff, novel treatment media are needed to sorb a broad range of energetic materials and to transform the retained compounds through abiotic and/or microbial processes.

Surface runoff of organic and inorganic contaminants from live-fire ranges is a challenging issue for the Department of Defense (DoD). Potentially even more problematic is the fact that inputs to surface waters from large testing and training ranges typically originate from multiple sources, often encompassing hundreds of acres. No available technologies are currently considered effective for controlling non-point source energetics-laden surface runoff. While numerous technologies exist to treat collected explosives residues, contaminated soil and even groundwater, the decentralized nature and sheer volume of military range runoff have precluded the use of treatment technologies at full scale in the field.

Range Runoff Treatment Technology Components

Based on the conceptual foundation of previous research into surface water runoff treatment for other contaminants, with a goal to “trap and treat” the target compounds, the following components were selected for inclusion in the technology developed to address range runoff contaminated with energetic compounds.

Peat

Previous research demonstrated that a peat-based system provided a natural and sustainable sorptive medium for organic explosives such as HMX, RDX, and TNT, allowing much longer residence times than predicted from hydraulic loading alone[12][13][14][15][16]. Peat moss represents a bioactive environment for treatment of the target contaminants. While the majority of the microbial reactions are aerobic due to the presence of measurable dissolved oxygen in the bulk solution, anaerobic reactions (including methanogenesis) can occur in microsites within the peat. The peat-based substrate acts not only as a long term electron donor as it degrades but also acts as a strong sorbent. This is important in intermittently loaded systems in which a large initial pulse of MC can be temporarily retarded on the peat matrix and then slowly degraded as they desorb[14][16]. This increased residence time enhances the biotransformation of energetics and promotes the immobilization and further degradation of breakdown products. Abiotic degradation reactions are also likely enhanced by association with the organic-rich peat (e.g., via electron shuttling reactions of humics)[17].

Soybean Oil

Modeling has indicated that peat moss amended with crude soybean oil would significantly reduce the flux of dissolved TNT, RDX, and HMX through the vadose zone to groundwater compared to a non-treated soil (see ESTCP ER-200434). The technology was validated in field soil plots, showing a greater than 500-fold reduction in the flux of dissolved RDX from macroscale Composition B detonation residues compared to a non-treated control plot[14]. Laboratory testing and modeling indicated that the addition of soybean oil increased the biotransformation rates of RDX and HMX at least 10-fold compared to rates observed with peat moss alone[16]. Subsequent experiments also demonstrated the effectiveness of the amended peat moss material for stimulating perchlorate transformation when added to a highly contaminated soil (Fuller et al., unpublished data). These previous findings clearly demonstrate the effectiveness of peat-based materials for mitigating transport of both organic and inorganic energetic compounds through soil to groundwater.

Biochar

Recent reports have highlighted additional materials that, either alone, or in combination with electron donors such as peat moss and soybean oil, may further enhance the sorption and degradation of surface runoff contaminants, including both legacy energetics and insensitive high explosives (IHE). For instance, biochar, a type of black carbon, has been shown to not only sorb a wide range of organic and inorganic contaminants including MCs[18][19][20][21], but also to facilitate their degradation[22][23][24][25][26][27]. Depending on the source biomass and pyrolysis conditions, biochar can possess a high specific surface area (on the order of several hundred m2/g)[28][29] and hence a high sorption capacity. Biochar and other black carbon also exhibit especially high affinity for nitroaromatic compounds (NACs) including TNT and 2,4-dinitrotoluene (DNT)[30][31][32]. This is due to the strong π-π electron donor-acceptor interactions between electron-rich graphitic domains in black carbon and the electron-deficient aromatic ring of the NAC

PFAS Screening Levels for Human Health and Ecological Risk Assessments

Human Health Screening Levels

Human health screening levels for PFAS have been modified multiple times over the last decade and, in the United States, are currently available for drinking water and soil exposures as Maximum Contaminant Levels (MCLs) and USEPA Regional Screening Levels (RSLs). USEPA finalized a National Primary Drinking Water Regulation (NPDWR) for six PFAS[33]:

  • Perfluorooctanoic acid (PFOA)
  • Perfluorooctane sulfonic acid (PFOS)
  • Perfluorohexane sulfonic acid (PFHxS)
  • Perfluorononanoic acid (PFNA)
  • Hexafluoropropylene oxide dimer acid (HFPO-DA, commonly known as GenX chemicals)
  • Perfluorobutane sulfonic acid (PFBS)

MCLs are enforceable drinking water standards based on the most recently available toxicity information that consider available treatment technologies and costs. The MCLs for PFAS include a Hazard Index of 1 for combined exposures to four PFAS. RSLs are developed for use in risk assessments and include soil and tap water screening levels for multiple PFAS. Soil RSLs are based on residential/unrestricted and commercial/industrial land uses, and calculations of site-specific RSLs are available.

Internationally, Canada and the European Union have also promulgated drinking water standards for select PFAS. However, large discrepancies exist among the various regulatory organizations, largely due to the different effect endpoints and exposure doses being used to calculate risk-based levels. The PFAS guidance from the Interstate Technology and Regulatory Council (ITRC) in the US includes a regularly updated compilation of screening values for PFAS and is available on their PFAS website[34]: https://pfas-1.itrcweb.org.

Ecological Screening Levels

Most peer-reviewed literature and regulatory-based environmental quality benchmarks have been developed using data for PFOS and PFOA; however, other select PFAAs have been evaluated for potential effects to aquatic receptors[34][35][36]. USEPA has developed water quality criteria for aquatic life[37][38][39] for PFOA and PFOS. Following extensive reviews of the peer-reviewed literature, Zodrow et al.[35] used the USEPA Great Lakes Initiative methodology[40] to calculate acute and chronic screening levels for aquatic life for 23 PFAS. The Argonne National Laboratory has also developed Ecological Screening Levels for multiple PFAS[41]. In contrast to surface water aquatic life benchmarks, sediment benchmark values are limited. For terrestrial systems, screening levels for direct exposure of soil plants and invertebrates to PFAS in soils have been developed for multiple AFFF-related PFAS[36][35], and the Canadian Council of Ministers of Environment developed several draft thresholds protective of direct toxicity of PFOS in soil[42].

Wildlife screening levels for abiotic media are back-calculated from food web models developed for representative receptors. Both Zodrow et al.[35] and Grippo et al.[41] include the development of risk-based screening levels for wildlife. The Michigan Department of Community Health[43] derived a provisional PFOS surface water value for avian and mammalian wildlife. In California, the San Francisco Bay Regional Water Quality Control Board developed terrestrial habitat soil ecological screening levels based on values developed in Zodrow et al.[35]. For PFOS only, a dietary screening level (i.e. applicable to the concentration of PFAS measured in dietary items) has been developed for mammals at 4.6 micrograms per kilogram (μg/kg) wet weight (ww), and for avians at 8.2 μg/kg ww[44].

Approaches for Evaluating Exposures and Effects in AFFF Site Environmental Risk Assessment: Human Health

Exposure pathways and effects for select PFAS are well understood, such that standard human health risk assessment approaches can be used to quantify risks for populations relevant to a site. Human health exposures via drinking water have been the focus in risk assessments and investigations at PFAS sites[45][46]. Risk assessment approaches for PFAS in drinking water follow typical, well-established drinking water risk assessment approaches for chemicals as detailed in regulatory guidance documents for various jurisdictions.

Incidental exposures to soil and dusts for PFAS can occur during a variety of soil disturbance activities, such as gardening and digging, hand-to-mouth activities, and intrusive groundwork by industrial or construction workers. As detailed by the ITRC[34], many US states and USEPA have calculated risk-based screening levels for these soil and drinking water pathways (and many also include dermal exposures to soils) using well-established risk assessment guidance.

Field and laboratory studies have shown that some PFCAs and PFSAs bioaccumulate in fish and other aquatic life at rates that could result in relevant dietary PFAS exposures for consumers of fish and other seafood[47][48][49][50][51][52][53][54][55][56]. In addition to fish, terrestrial wildlife can accumulate contaminants from impacted sites, resulting in potential exposures to consumers of wild game[57]. Additionally, exposures can occur though consumption of homegrown produce or agricultural products that originate from areas irrigated with PFAS-impacted groundwater, or that are amended with biosolids that contain PFAS, or that contain soils that were directly affected by PFAS releases[58]. Multiple studies have found PFAS can be taken up by plants from soil porewater[59][60][61], and livestock can accumulate PFAS from drinking water and/or feed[62]. Thus, when PFAS are present in surface water bodies where fishing or shellfish harvesting occurs or terrestrial areas where produce is grown or game is hunted, the bioaccumulation of PFAS into dietary items can be an important pathway for human exposure.

PFAAs such as PFOA and PFOS are not expected to volatilize from PFAS-impacted environmental media[63][64] such as soil and groundwater, which are the primary focus of most site-specific risk assessments. In contrast to non-volatile PFAAs, fluorotelomer alcohols (FTOHs) are among the more widely studied of the volatile PFAS. FTOHs are transient in the atmosphere with a lifetime of 20 days[65]. At most AFFF sites under evaluation, AFFF releases have occurred many years before such that FTOH may no longer be present. As such, the current assumption is that volatile PFAS, such as FTOHs historically released at the site, will have transformed to stable, low-volatility PFAS, such as PFAAs in soil or groundwater, or will they have diffused to the outdoor atmosphere. There is no evidence that FTOHs or other volatile PFAS are persistent in groundwater or soils such that they present an indoor vapor intrusion pathway risk concern as observed for chlorinated solvents. Ongoing research continues for the vapor pathway[34].

General and site-specific human health exposure pathways and risk assessment methods as outlined by USEPA[66][67] can be applied to PFAS risk assessments for which human health toxicity values have been developed. Additionally, for risk assessments with dietary exposures of PFAS, standard risk assessment food web modeling can be used to develop initial estimates of dietary concentrations which can be confirmed with site-specific tissue sampling programs.

Approaches for Evaluating Exposures and Effects in AFFF Site Environmental Risk Assessment: Ecological

Information available currently on exposures and effects of PFAS in ecological receptors indicate that the PFAS ecological risk issues at most sites are primarily associated with risks to vertebrate wildlife. Avian and mammalian wildlife are relatively sensitive to PFAS, and dietary intake via bioaccumulation in terrestrial and aquatic food webs can result in exposures that are dominated by the more accumulative PFAS[68][36][35]. Direct toxicity to aquatic life (e.g., fish, pelagic life, benthic invertebrates, and aquatic plants) can occur from exposure to sediment and surface water at effected sites. For larger areas, surface water concentrations associated with adverse effects to aquatic life are generally higher than those that could result in adverse effects to aquatic-dependent wildlife. Soil invertebrates and plants are generally less sensitive, with risk-based concentrations in soil being much higher than those associated with potential effects to terrestrial wildlife[35].

Aquatic life are exposed to PFAS through direct exposure in surface water and sediment. Ecological risk assessment approaches for PFAS for aquatic life follow standard risk assessment approaches. The evaluation of potential risks for aquatic life with direct exposure to PFAS in environmental media relies on comparing concentrations in external exposure media to protective, media-specific benchmarks, including the aquatic life risk-based screening levels discussed above[35][69].

When an area at the point of PFAS release is an industrial setting which does not feature favorable habitats for terrestrial and aquatic-dependent wildlife, the transport mechanisms may allow PFAS to travel offsite. If offsite or downgradient areas contain ecological habitat, then PFAS transported to these areas are expected to pose the highest risk potential to wildlife, particularly those areas that feature aquatic habitat[70][68].

Wildlife receptors, specifically birds and mammals, are typically exposed to PFAS through uptake from dietary sources such as plants and invertebrates, along with direct soil ingestion during foraging activities. Dietary intake modeling typical for ecological risk assessments is the recommended approach for an evaluation of potential risks to wildlife species where PFAS exposure occurs primarily via dietary uptake from bioaccumulation pathways. Dietary intake modeling uses relevant exposure factors for each receptor group (terrestrial birds, terrestrial mammals, aquatic-dependent birds, and aquatic mammals) to determine a total daily intake (TDI) of PFAS via all potential exposure pathways. This approach requires determination of concentrations of PFAS in dietary items, which can be obtained by measuring PFAS in biota at sites or by using food web models to predict concentrations in biota using measured concentrations of PFAS in soil, sediment, or surface water. Food web models use bioaccumulation metrics such as bioaccumulation factors (BAFs) and biomagnification factors (BMFs) with measurements of PFAS in abiotic media to estimate concentrations in dietary items, including plants and benthic or pelagic invertebrates, to model wildlife exposure and calculate TDI. Once site-specific TDI values are calculated, they are compared to known TRVs identified from toxicity data with exposure doses associated with a lack of adverse effects (termed no observed adverse effect level [NOAEL]) or low adverse effects (termed lowest observed adverse effect level [LOAEL]), per standard risk assessment practice[67].

Recently, Conder et al.[36], Gobas et al.[71], and Zodrow et al.[35] compiled bioaccumulation modeling parameters and approaches for terrestrial and aquatic food web modeling of a variety of commonly detected PFAS at AFFF sites. There are also several sources of TRVs which can be relied upon for estimating TDI values[36][71][35][72][73]. In general, the highest risk for PFAS is expected for smaller insectivore and omnivore receptors (e.g., shrews and other small rodents, small nonmigratory birds), which tend to be lower in trophic level and spend more time foraging in small areas similar to or smaller in size than the impacted area. Compared to smaller, lower-trophic level organisms, larger mammalian and avian carnivores are expected to have lower exposures from site-specific PFAS sources because they forage over larger areas that may include areas that are not impacted, as compared to small organisms with small home ranges[68][36][71][73][35].

Available information regarding PFAS exposure pathways and effects in aquatic life, terrestrial invertebrates and plants, as well as aquatic and terrestrial wildlife allow ecological risk assessment methods to be applied as outlined by USEPA[67] to site-specific PFAS risk assessments. Additionally, food web modeling can be used in site-specific PFAS risk assessment to develop initial estimates of dietary concentrations for aquatic and terrestrial wildlife, which can be confirmed with tissue sampling programs at a site.

PFAS Risk Assessment Data Gaps

There are a number of data gaps currently associated with PFAS risk assessment including the following:

  • Unmeasured PFAS: There are a number of additional PFAS that we know little about and many PFAS that we are unable to quantify in the environment. The approach to dealing with the lack of information on the overwhelming number of PFAS is being debated; in the meantime, however, PFAS beyond PFOS and PFOA are being studied more, and this information will result in improved characterization of risks for other PFAS.
  • Mixtures: Another major challenge in effects assessment for PFAS, for both human health risk assessments and environmental risk assessments, is understanding the potential importance of mixtures of PFAS. Considering the limited human health and ecological toxicity data available for just a few PFAS, the understanding of the relative toxicity, additivity, or synergistic effects of PFAS in mixtures is just beginning.
  • Toxicity Data Gaps: For environmental risk assessments, some organisms such as reptiles and benthic invertebrates do not have toxicity data available. Benchmark or threshold concentrations of PFAS in environmental media intended to be protective of wildlife and aquatic organisms suffer from significant uncertainty in their derivation due to the limited number of species for which data are available. As species-specific data becomes available for more types of organisms, the accuracy of environmental risk assessments is likely to improve.

References

  1. ^ Sansalone, J.J., 1999. In-situ performance of a passive treatment system for metal source control. Water Science and Technology, 39(2), pp. 193-200. doi: 10.1016/S0273-1223(99)00023-2
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See Also

Agency for Toxic Substances and Disease Registry (ATSDR) PFAS Health Studies

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