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==1,2,3-Trichloropropane (TCP)==
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==Remediation of Stormwater Runoff Contaminated by Munition Constituents==  
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[[Wikipedia: 1,2,3-Trichloropropane | 1,2,3-Trichloropropane (TCP)]] is a chlorinated volatile organic compound (CVOC) that has been used in chemical production processes, in agriculture, and as a solvent, resulting in point and non-point source contamination of soil and groundwater.  TCP is mobile and highly persistent in soil and groundwater. TCP is not currently regulated at the national level in the United States, but [[Wikipedia: Maximum contaminant level | maximum contaminant levels (MCLs)]] have been developed by some states.  Current treatment methods for TCP are limited and can be cost prohibitive. However, some treatment approaches, particularly [[Chemical Reduction (In Situ - ISCR) | ''in situ'' chemical reduction (ISCR)]] with [[Wikipedia: In_situ_chemical_reduction#Zero_valent_metals_%28ZVMs%29 | zero valent zinc (ZVZ)]] and [[Bioremediation - Anaerobic | ''in situ'' bioremediation (ISB)]], have recently been shown to have potential as practical remedies for TCP contamination of groundwater.
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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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*[[Bioremediation - Anaerobic | Anaerobic Bioremediation]]
 
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*[[Chemical Reduction (In Situ - ISCR) | ''In Situ'' Chemical Reduction (ISCR)]]
 
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*[[Chemical Oxidation (In Situ - ISCO) | ''In Situ'' Chemical Oxidation (ISCO)]]
 
  
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'''Contributor(s):'''  
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*[[Munitions Constituents]]
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*[[Dr. Alexandra Salter-Blanc | Alexandra J. Salter-Blanc]]
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*[[Dr. Paul Tratnyek | Paul G. Tratnyek]]
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*John Merrill
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'''Contributor:''' Mark E. Fuller
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*Alyssa Saito
 
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*Lea Kane
 
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*Eric Suchomel
 
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*[[Dr. Rula Deeb | Rula Deeb]]
 
  
 
'''Key Resource(s):'''
 
'''Key Resource(s):'''
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*Prospects for Remediation of 1,2,3-Trichloropropane by Natural and Engineered Abiotic Degradation Reactions. Strategic Environmental Research and Development Program (SERDP), Project ER-1457.<ref name="Tratnyek2010">Tratnyek, P.G., Sarathy, V., Salter, A.J., Nurmi, J.T., O’Brien Johnson, G., DeVoe, T., and Lee, P., 2010. Prospects for Remediation of 1,2,3-Trichloropropane by Natural and Engineered Abiotic Degradation Reactions. Strategic Environmental Research and Development Program (SERDP), Project ER-1457. [https://serdp-estcp.org/Program-Areas/Environmental-Restoration/Contaminated-Groundwater/Emerging-Issues/ER-1457/ER-1457/(language)/eng-US  Website]&nbsp;&nbsp; [[Media: ER-1457-FR.pdf | Report.pdf]]</ref>
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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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*Verification Monitoring for In Situ Chemical Reduction Using Zero-Valent Zinc, A Novel Technology for Remediation of Chlorinated Alkanes. Strategic Environmental Research and Development Program (SERDP), Project ER-201628.<ref name="Kane2020">Kane, L.Z., Suchomel, E.J., and Deeb, R.A., 2020. Verification Monitoring for In Situ Chemical Reduction Using Zero-Valent Zinc, A Novel Technology for Remediation of Chlorinated Alkanes. Strategic Environmental Research and Development Program (SERDP), Project ER-201628. [https://www.serdp-estcp.org/Program-Areas/Environmental-Restoration/Contaminated-Groundwater/Persistent-Contamination/ER-201628  Website]&nbsp;&nbsp; [[Media: ER-201628.pdf | Report.pdf]]</ref>
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==Background==
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===Surface Runoff Characteristics and Treatment Approaches===
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[[File: FullerFig1.png | thumb | 300 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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==Introduction==
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===Surface Runoff on Ranges===
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[[File:123TCPFig1.png|thumb|left|Figure 1. Ball and stick representation of the molecular structure of TCP (Salter-Blanc and Tratnyek, unpublished)]]
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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>.
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1,2,3-Trichloropropane (TCP) (Figure 1) is a man-made chemical that was used in the past primarily as a solvent and extractive agent, as a paint and varnish remover, and as a cleaning and degreasing agent.<ref name="ATSDR2021"> Agency for Toxic Substances and Disease Registry (ATSDR), 2021. Toxicological Profile for 1,2,3-Trichloropropane. Free download from: [https://www.atsdr.cdc.gov/toxprofiles/tp57.pdf ATSDR]&nbsp;&nbsp; [[Media: TCP2021ATSDR.pdf | Report.pdf]]</ref>. Currently, TCP is primarily used in chemical synthesis of compounds such as [[Wikipedia: Polysulfone | polysulfone]] liquid polymers used in the aerospace and automotive industries; [[Wikipedia: Hexafluoropropylene | hexafluoropropylene]] used in the agricultural, electronic, and pharmaceutical industries; [[Wikipedia: Polysulfide | polysulfide]] polymers used as sealants in manufacturing and construction; and [[Wikipedia: 1,3-Dichloropropene | 1,3-dichloropropene]] used in agriculture as a soil fumigant. TCP may also be present in products containing these chemicals as an impurity<ref name="ATSDR2021"/><ref name="CH2M2005">CH2M HILL, 2005. Interim Guidance for Investigating Potential 1,2,3-Trichloropropane Sources in San Gabriel Valley Area 3. [[Media: INTERIM_GUIDANCE_FOR_INVESTIGATING_POTENTIAL_1%2C2%2C3-TRICHLOROPROPANE_SOURCES.pdf | Report.pdf]]&nbsp;&nbsp;  [https://cumulis.epa.gov/supercpad/cursites/csitinfo.cfm?id=0902093  Website]</ref>. For example, the 1,2-dichlropropane/1,3-dichloropropene soil fumigant mixture (trade name D-D), which is no longer sold in the United States, contained TCP as an impurity and has been linked to TCP contamination in groundwater<ref name="OkiGiambelluca1987">Oki, D.S. and Giambelluca, T.W., 1987. DBCP, EDB, and TCP Contamination of Ground Water in Hawaii. Groundwater, 25(6), pp. 693-702.  [https://doi.org/10.1111/j.1745-6584.1987.tb02210.x DOI: 10.1111/j.1745-6584.1987.tb02210.x]</ref><ref name="CH2M2005"/>. Soil fumigants currently in use which are composed primarily of 1,3-dichloropropene may also contain TCP as an impurity, for instance Telone II has been reported to contain up to 0.17 percent TCP by weight<ref name="Kielhorn2003">Kielhorn, J., Könnecker, G., Pohlenz-Michel, C., Schmidt, S. and Mangelsdorf, I., 2003. Concise International Chemical Assessment Document 56: 1,2,3-Trichloropropane. World Health Organization, Geneva. [http://www.who.int/ipcs/publications/cicad/en/cicad56.pdf Website]&nbsp;&nbsp; [[Media: WHOcicad56TCP.pdf | Report.pdf]]</ref>.  
 
  
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TCP contamination is problematic because it is “reasonably anticipated to be a human carcinogen” based on evidence of carcinogenicity to animals<ref name="NTP2016"> National Toxicology Program, 2016. Report on Carcinogens, 14th ed. U.S. Department of Health and Human Services, Public Health Service. Free download from: [https://ntp.niehs.nih.gov/ntp/roc/content/profiles/trichloropropane.pdf  NIH]&nbsp;&nbsp; [[Media: NTP2016trichloropropane.pdf | Report.pdf]]</ref>. Toxicity to humans appears to be high relative to other chlorinated solvents<ref name="Kielhorn2003"/>, suggesting that even low-level exposure to TCP could pose a significant human health risk.
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==Toxicological Effects of PFAS==
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The characterization of toxicological effects in human health risk assessments is based on toxicological studies of mammalian exposures to per- and polyfluoroalkyl substances (PFAS), primarily studies involving [[Wikipedia:Perfluorooctanesulfonic acid | perfluorooctanesulfonic acid (PFOS)]] and [[Wikipedia:Perfluorooctanoic acid|perfluorooctanoic acid (PFOA)]]. The most sensitive noncancer adverse effects involve the liver and kidney, immune system, and various developmental and reproductive endpoints<ref name="USEPA2024b">United States Environmental Protection Agency (USEPA), 2024. Per- and Polyfluoroalkyl Substances (PFAS) Final PFAS National Primary Drinking Water Regulation. [https://www.epa.gov/sdwa/and-polyfluoroalkyl-substances-pfas Website]</ref>. A select number of PFAS have been evaluated for carcinogenicity, primarily using epidemiological data. Only PFOS and PFOA (and their derivatives) have sufficient data for USEPA to characterize as ''Likely to Be Carcinogenic to Humans'' via the oral route of exposure. Epidemiological studies provided evidence of bladder, prostate, liver, kidney, and breast cancers in humans related to PFOS exposure, as well as kidney and testicular cancer in humans and limited evidence of breast cancer related to PFOA exposure<ref name="USEPA2024b"/><ref name="USEPA2016a">United States Environmental Protection Agency (USEPA), 2016. Drinking Water Health Advisory for Perfluorooctane Sulfonate (PFOS). Office of Water, EPA 822-R-16-004. [https://www.epa.gov/sites/production/files/2016-05/documents/pfos_health_advisory_final-plain.pdf  Free Download]&nbsp; [[Media: USEPA-2016-pfos_health_advisory_final-plain.pdf | Report.pdf]]</ref><ref name="USEPA2016b">United States Environmental Protection Agency (USEPA), 2016b. Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA). Office of Water, EPA 822-R-16-005. [https://www.epa.gov/sites/production/files/2016-05/documents/pfoa_health_advisory_final_508.pdf Free Download]&nbsp; [[Media: pfoa_EPA 822-R-16-005.pdf | Report.pdf]]</ref>.
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USEPA’s Integrated Risk Management System (IRIS) Program is developing Toxicological Reviews to improve understanding of the toxicity of several additional PFAS (i.e., not solely PFOA and PFOS). Toxicological Reviews provide an overview of cancer and noncancer health effects based on current literature and, where data are sufficient, derive human health toxicity criteria (i.e., human health oral reference doses and cancer slope factors) that form the basis for risk-based decision making. For risk assessors, these documents provide USEPA reference doses and cancer slope factors that can be used with exposure information and other considerations to assess human health risk. Final Toxicological Reviews have been completed for the following PFAS:
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*Perfluorooctanesulfonic acid (PFOS)
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*Perfluorooctanoic acid (PFOA)
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*Perfluorobutanoic acid (PFBA)
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*Perfluorohexanoic acid (PFHxA)
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*Perfluorobutane sulfonic acid (PFBS)
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*Perfluoropropionic acid (PFPrA)
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*Perfluorohexane sulfonic acid (PFHxS)
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*Lithium bis[(trifluoromethyl)sulfonyl]azanide (HQ-115)
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*Hexafluoropropylene oxide dimer acid (HFPO DA) and its Ammonium Salt
  
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==Environmental Fate==
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Toxicity assessments are ongoing for the following PFAS:
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TCP’s fate in the environment is governed by its physical and chemical properties (Table 1). TCP does not adsorb strongly to soil, making it likely to leach into groundwater and exhibit high mobility. In addition, TCP is moderately volatile and can partition from surface water and moist soil into the atmosphere. Because TCP is only slightly soluble and denser than water, it can form a [[Wikipedia: Dense non-aqueous phase liquid | dense non-aqueous phase liquid (DNAPL)]] as observed at the Tyson’s Dump Superfund Site<ref name="USEPA2019"> United States Environmental Protection Agency (USEPA), 2019. Fifth Five-year Review Report, Tyson’s Dump Superfund Site, Upper Merion Township, Montgomery County, Pennsylvania. Free download from: [https://semspub.epa.gov/work/03/2282817.pdf USEPA]&nbsp;&nbsp; [[Media: USEPA2019.pdf | Report.pdf]]</ref>. TCP is generally resistant to aerobic biodegradation, hydrolysis, oxidation, and reduction under naturally occurring conditions making it persistent in the environment<ref name="Tratnyek2010"/>.
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*Perfluorononanoic acid (PFNA)
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*Perfluorodecanoic acid (PFDA)  
  
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{| class="wikitable" style="float:right; margin-left:10px;text-align:center;"
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It is important to note human health toxicity criteria for inhalation of PFAS are not included in the Final Toxicological Reviews and are not currently available.  
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|+Table 1.  Physical and chemical properties of TCP<ref name="USEPA2017">United States Environmental Protection Agency (USEPA), 2017. Technical Fact Sheet—1,2,3-Trichloropropane (TCP). EPA Project 505-F-17-007. 6 pp.  Free download from: [https://www.epa.gov/sites/production/files/2017-10/documents/ffrrofactsheet_contaminants_tcp_9-15-17_508.pdf  USEPA]&nbsp;&nbsp; [[Media: epa_tcp_2017.pdf | Report.pdf]]</ref>
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In addition to IRIS, state agencies have developed peer-reviewed provisional toxicity values that have been incorporated into USEPA’s RSLs, which are updated biannually. These values have not been reviewed by or incorporated into IRIS.  
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|-
 
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!Property
 
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!Value
 
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| Chemical Abstracts Service (CAS) Number || 96-18-4
 
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| Physical Description</br>(at room temperature) || Colorless to straw-colored liquid
 
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|-
 
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| Molecular weight</br>(g/mol) || 147.43
 
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|-
 
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| Water solubility at 25°C</br>(mg/L)|| 1,750 (slightly soluble)
 
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|-
 
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| Melting point</br>(°C)|| -14.7
 
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|-
 
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| Boiling point</br>(°C) || 156.8
 
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|-
 
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| Vapor pressure at 25°C</br>(mm Hg) || 3.10 to 3.69
 
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|-
 
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| Density at 20°C (g/cm<sup>3</sup>) || 1.3889
 
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|-
 
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| Octanol-water partition coefficient</br>(log''K<sub>ow</sub>'') || 1.98 to 2.27</br>(temperature dependent)
 
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|-
 
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| Organic carbon-water partition coefficient</br>(log''K<sub>oc</sub>'') || 1.70 to 1.99</br>(temperature dependent)
 
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| Henry’s Law constant at 25°C</br>(atm-m<sup>3</sup>/mol) || 3.17x10<sup>-4</sup><ref name="ATSDR2021"/> to 3.43x10<sup>-4</sup><ref name="LeightonCalo1981">Leighton Jr, D.T. and Calo, J.M., 1981. Distribution Coefficients of Chlorinated Hydrocarbons in Dilute Air-Water Systems for Groundwater Contamination Applications. Journal of Chemical and Engineering Data, 26(4), pp. 382-385.  [https://doi.org/10.1021/je00026a010 DOI: 10.1021/je00026a010]</ref>
 
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|}
 
  
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==Occurrence==
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With respect to ecological toxicity, effects on reproduction, growth, and development of avian and mammalian wildlife have been documented in controlled laboratory studies of exposures of standard toxicological test species (e.g., mice, quail) to PFAS. Many of these studies have been reviewed<ref name="ConderEtAl2020"> Conder, J., Arblaster, J., Larson, E., Brown, J., Higgins, C., 2020. Guidance for Assessing the Ecological Risks of PFAS to Threatened and Endangered Species at Aqueous Film Forming Foam-Impacted Sites. Strategic Environmental Research and Development Program (SERDP) Project ER 18-1614. [https://serdp-estcp.mil/projects/details/3f890c9b-7f72-4303-8d2e-52a89613b5f6 Project Website]&nbsp; [[Media: ER18-1614_Guidance.pdf | Guidance Document]]</ref><ref name="GobasEtAl2020">Gobas, F.A.P.C., Kelly, B.C., Kim, J.J., 2020. Final Report: A Framework for Assessing Bioaccumulation and Exposure Risks of PFAS in Threatened and Endangered Species on AFFF-Impacted Sites. SERDP Project ER18-1502. [https://serdp-estcp.mil/projects/details/09c93894-bc73-404a-8282-51196c4be163 Project Website]&nbsp; [[Media: ER18-1502_Final.pdf | Final Report]]</ref><ref name="Suski2020">Suski, J.G., 2020. Investigating Potential Risk to Threatened and Endangered Species from Per- and Polyfluoroalkyl Substances (PFAS) on Department of Defense (DoD) Sites. SERDP Project ER18-1626. [https://serdp-estcp.mil/projects/details/c328f8e3-95a4-4820-a0d4-ef5835134636 Project Website]&nbsp; [[Media: ER18-1626_Final.pdf | Report.pdf]]</ref><ref name="ZodrowEtAl2021a">Zodrow, J.M., Frenchmeyer, M., Dally, K., Osborn, E., Anderson, P. and Divine, C., 2021. Development of Per and Polyfluoroalkyl Substances Ecological Risk-Based Screening Levels. Environmental Toxicology and Chemistry, 40(3), pp. 921-936. [https://doi.org/10.1002/etc.4975 doi: 10.1002/etc.4975]&nbsp;&nbsp; [[Media: ZodrowEtAl2021a.pdf | Open Access Article]]</ref> to derive ecological Toxicity Reference Values (TRVs). TRVs can be used alongside exposure information and other considerations to assess ecological risk. Avian and mammalian wildlife receptors are generally expected to have the highest risks due to PFAS exposure. Direct toxicity to aquatic life, such as fish and invertebrates, from exposure to sediment and surface water also occurs, though concentrations in water 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 less sensitive to PFAS when compared to terrestrial wildlife, with risk-based PFAS concentrations in soil being much higher than those associated with potential effects to terrestrial wildlife<ref name="ZodrowEtAl2021a"/>.
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TCP has been detected in approximately 1% of public water supply and domestic well samples tested by the United States Geological Survey. More specifically, TCP was detected in 1.2% of public supply well samples collected between 1993 and 2007 by Toccalino and Hopple<ref name="ToccalinoHopple2010">Toccalino, P.L., Norman, J.E., Hitt, K.J., 2010. Quality of Source Water from Public-Supply Wells in the United States, 1993–2007. Scientific Investigations Report 2010-5024. U.S. Geological Survey. [https://doi.org/10.3133/sir20105024 DOI: 10.3133/sir20105024]  Free download from: [https://pubs.er.usgs.gov/publication/sir20105024 USGS]&nbsp;&nbsp; [[Media: Quality_of_source_water_from_public-supply_wells_in_the_United_States%2C_1993-2007.pdf | Report.pdf]]</ref> and 0.66% of domestic supply well samples collected between 1991 and 2004 by DeSimone<ref name="DeSimone2009">DeSimone, L.A., 2009. Quality of Water from Domestic Wells in Principal Aquifers of the United States, 1991–2004. U.S. Geological Survey, Scientific Investigations Report 2008–5227. 139 pp. Free download from: [http://pubs.usgs.gov/sir/2008/5227 USGS]&nbsp;&nbsp; [[Media: DeSimone2009.pdf | Report.pdf]]</ref>. TCP was detected at a higher rate in domestic supply well samples associated with agricultural land-use studies than samples associated with studies comparing primary aquifers (3.5% versus 0.2%)<ref name="DeSimone2009"/>.  
 
  
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==Regulation==
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==PFAS Screening Levels for Human Health and Ecological Risk Assessments==
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The United States Environmental Protection Agency (USEPA) has not established an MCL for TCP, although guidelines and health standards are in place<ref name="USEPA2017"/>. TCP was included in the Contaminant Candidate List 3<ref name="USEPA2009">United States Environmental Protection Agency (US EPA), 2009. Drinking Water Contaminant Candidate List 3-Final. Federal Register 74(194), pp. 51850–51862, Document E9-24287. [https://www.federalregister.gov/documents/2009/10/08/E9-24287/drinking-water-contaminant-candidate-list-3-final Website]&nbsp;&nbsp; [[Media: FR74-194DWCCL3.pdf | Report.pdf]]</ref> and the Unregulated Contaminant Monitoring Rule 3 (UCMR 3)<ref name="USEPA2012">United States Environmental Protection Agency (US EPA), 2012. Revisions to the Unregulated Contaminant Mentoring Regulation (UCMR 3) for Public Water Systems. Federal Register 77(85) pp. 26072-26101. [https://www.federalregister.gov/documents/2012/05/02/2012-9978/revisions-to-the-unregulated-contaminant-monitoring-regulation-ucmr-3-for-public-water-systems  Website]&nbsp;&nbsp; [[Media: FR77-85UCMR3.pdf | Report.pdf]]</ref>. The UCMR 3 specified that data be collected on TCP occurrence in public water systems over the period of January 2013 through December 2015 against a reference concentration range of 0.0004 to 0.04 μg/L<ref name="USEPA2017a">United States Environmental Protection Agency (USEPA), 2017. The Third Unregulated Contaminant Monitoring Rule (UCMR 3): Data Summary. EPA 815-S-17-001. [https://www.epa.gov/dwucmr/data-summary-third-unregulated-contaminant-monitoring-rule  Website]&nbsp;&nbsp; [[Media: ucmr3-data-summary-january-2017.pdf | Report.pdf]]</ref>. The reference concentration range was determined based on a cancer risk of 10-6 to 10-4 and derived from an oral slope factor of 30 mg/kg-day, which was determined by the EPA’s Integrated Risk Information System<ref name="IRIS2009">USEPA Integrated Risk Information System (IRIS), 2009. 1,2,3-Trichloropropane (CASRN 96-18-4). [https://cfpub.epa.gov/ncea/iris2/chemicalLanding.cfm?substance_nmbr=200 Website]&nbsp;&nbsp; [[Media: TCPsummaryIRIS.pdf | Summary.pdf]]</ref>. Of 36,848 samples collected during UCMR 3, 0.67% exceeded the minimum reporting level of 0.03 µg/L. 1.4% of public water systems had at least one detection over the minimum reporting level, corresponding to 2.5% of the population<ref name="USEPA2017a"/>. While these occurrence percentages are relatively low, the minimum reporting level of 0.03 µg/L is more than 75 times the USEPA-calculated Health Reference Level of 0.0004 µg/L. Because of this, TCP may occur in public water systems at concentrations that exceed the Health Reference Level but are below the minimum reporting level used during UCMR 3 data collection. These analytical limitations and lack of lower-level occurrence data have prevented the USEPA from making a preliminary regulatory determination for TCP<ref name="USEPA2021">USEPA, 2021. Announcement of Final Regulatory Determinations for Contaminants on the Fourth Drinking Water Contaminant Candidate List. Free download from: [https://www.epa.gov/sites/default/files/2021-01/documents/10019.70.ow_ccl_reg_det_4.final_web.pdf USEPA]&nbsp;&nbsp; [[Media: CCL4.pdf | Report.pdf]]</ref>.
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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)
 +
*Perfluorooctane sulfonic acid (PFOS)
 +
*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)
  
−
Some US states have established their own standards including Hawaii which has established an MCL of 0.6 μg/L<ref name="HDOH2013">Hawaii Department of Health, 2013. Amendment and Compilation of Chapter 11-20 Hawaii Administrative Rules. Free download from: [http://health.hawaii.gov/sdwb/files/2016/06/combodOPPPD.pdf Hawaii Department of Health]&nbsp;&nbsp; [[Media: Amendment_and_Compilation_of_Chapter_11-20_Hawaii_Administrative_Rules.pdf | Report.pdf]]</ref>. California has established an MCL of 0.005 μg/L<ref name="CCR2021">California Code of Regulations, 2021. Section 64444 Maximum Contaminant Levels – Organic Chemicals (22 CA ADC § 64444). [https://govt.westlaw.com/calregs/Document/IA7B3800D18654ABD9E2D24A445A66CB9 Website]</ref>,  a notification level of 0.005 μg/L, and a public health goal of 0.0007 μg/L<ref name="OEHHA2009">Office of Environmental Health Hazard Assessment (OEHHA), California Environmental Protection Agency, 2009. Final Public Health Goal for 1,2,3-Trichloropropane in Drinking Water. [https://oehha.ca.gov/water/public-health-goal/final-public-health-goal-123-trichloropropane-drinking-water Website]</ref>, and New Jersey has established an MCL of 0.03 μg/L<ref name="NJAC2020">New Jersey Administrative Code 7:10, 2020. Safe Drinking Water Act Rules. Free download from: [https://www.nj.gov/dep/rules/rules/njac7_10.pdf New Jersey Department of Environmental Protection]</ref>.
+
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.   
  
−
==Transformation Processes==
+
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.
  
 +
===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<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>.
  
−
{| class="wikitable" style="float:right; margin-left:10px;text-align:center;"
+
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>.
−
|+Table 2. Advantages and limitations of TCP treatment technologies
 
−
|-
 
−
! Technology
 
−
! Advantages
 
−
! Limitations
 
−
|-
 
−
| ZVZ
 
−
| style="text-align:left;" |
 
−
* Can degrade TCP at relatively high and low concentrations
 
−
* Faster reaction rates than ZVI
 
−
* Material is commercially available
 
−
| style="text-align:left;" |
 
−
* Higher cost than ZVI
 
−
* Difficult to distribute in subsurface ''in situ'' applications
 
−
|-
 
−
| Groundwater</br>Extraction and</br>Treatment
 
−
| style="text-align:left;" |
 
−
* Can cost-effectively capture and treat larger, more dilute</br>groundwater plumes than ''in situ'' technologies
 
−
* Well understood and widely applied technology
 
−
| style="text-align:left;" |
 
−
* Requires construction, operation and maintenance of</br>aboveground treatment infrastructure
 
−
* Typical technologies (e.g. GAC) may be expensive due</br>to treatment inefficiencies
 
−
|-
 
−
| ZVI
 
−
| style="text-align:left;" |
 
−
* Can degrade TCP at relatively high and low concentrations
 
−
* Lower cost than ZVZ
 
−
* Material is commercially available
 
−
| style="text-align:left;" |
 
−
* Lower reactivity than ZVZ, therefore may require higher</br>ZVI volumes or thicker PRBs
 
−
* Difficult to distribute in subsurface ''in situ'' applications
 
−
|-
 
−
| ISCO
 
−
| style="text-align:left;" |
 
−
* Can degrade TCP at relatively high and low concentrations
 
−
* Strategies to distribute amendments ''in situ'' are well established
 
−
* Material is commercially available
 
−
| style="text-align:left;" |
 
−
* Most effective oxidants (e.g., base-activated or heat-activated</br>persulfate) are complex to implement
 
−
* Secondary water quality impacts (e.g., high pH, sulfate, </br>hexavalent chromium) may limit ability to implement
 
−
|-
 
−
| ''In Situ''</br>Bioremediation
 
−
| style="text-align:left;" |
 
−
* Can degrade TCP at moderate to high concentrations
 
−
* Strategies to distribute amendments ''in situ'' are well established
 
−
* Materials are commercially available and inexpensive
 
−
| style="text-align:left;" |
 
−
* Slower reaction rates than ZVZ or ISCO
 
−
|}
 
  
 +
==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<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.
  
 +
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.
  
−
There&nbsp;are&nbsp;two&nbsp;main&nbsp;approaches to downscaling. One method, commonly referred to as “statistical downscaling”, uses the empirical-statistical relationships between large-scale weather phenomena and historical local weather data. In this method, these statistical relationships are applied to output generated by global climate models. A second method uses physics-based numerical models (regional-scale climate models or RCMs) of weather and climate that operate over a limited region of the earth (e.g., North America) and at spatial resolutions that are typically 3 to 10 times finer than the global-scale climate models. This method is known as “dynamical downscaling”. These regional-scale climate models are similar to the global models with respect to their reliance on the principles of physics, but because they operate over only part of the earth, they require information about what is coming in from the rest of the earth as well as what is going out of the limited region of the model. This is generally obtained from a global model. The primary differences between statistical and dynamical downscaling methods are summarized in Table 1.
+
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.  
  
−
It&nbsp;is&nbsp;important&nbsp;to&nbsp;realize that there is no “best” downscaling method or dataset, and that the best method/dataset for a given problem depends on that problem’s specific needs. Several data products based on downscaling higher level spatial data are available ([https://cida.usgs.gov/gdp/ USGS], [http://maca.northwestknowledge.net/ MACA], [https://www.narccap.ucar.edu/ NARCCAP], [https://na-cordex.org/ CORDEX-NA]). The appropriate method and dataset to use depends on the intended application. The method selected should be able to credibly resolve spatial and temporal scales relevant for the application. For example, to develop a risk analysis of frequent flooding, the data product chosen should include precipitation at greater than a diurnal frequency and over multi-decadal timescales. This kind of product is most likely to be available using the dynamical downscaling method. SERDP reviewed the various advantages and disadvantages of using each type of downscaling method and downscaling dataset, and developed a recommended process that is publicly available<ref name="Kotamarthi2016"/>. In general, the following recommendations should be considered in order to pick the right downscaled dataset for a given analysis:
+
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"/>.
  
−
* When a problem depends on using a large number of climate models and emission scenarios to perform preliminary assessments and to understand the uncertainty range of projections, then using a statistical downscaled dataset is recommended. 
+
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.
−
* When the assessment needs a more extensive parameter list or is analyzing a region with few long-term observational data, dynamically downscaled climate change projections are recommended.
 
  
−
==Uncertainty in Projections==
+
==Approaches for Evaluating Exposures and Effects in AFFF Site Environmental Risk Assessment: Ecological==
−
{| class="wikitable" style="float:right; margin-left:10px;text-align:center;"
+
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"/>.
−
|+Table 2.  Downscaling model characteristics and output<ref name="Kotamarthi2016"/>
 
−
|-
 
−
!Model or</br>Dataset Name
 
−
!Model<br />Method
 
−
!Output<br />Variables
 
−
!Output<br />Format
 
−
!Spatial</br>Resolution
 
−
!Time</br>Resolution
 
−
|-
 
−
| colspan="6" style="text-align: left; background-color:white;" |'''Statistical Downscaled Datasets'''
 
−
|-
 
−
| [https://worldclim.org/data/index.html WorldClim]<ref name="Hijmans2005">Hijmans, R.J., Cameron, S.E., Parra, J.L., Jones, P.G. and Jarvis, A., 2005. Very High Resolution Interpolated Climate Surfaces for Global Land Areas. International Journal of Climatology: A Journal of the Royal Meteorological Society, 25(15), pp 1965-1978.  [https://doi.org/10.1002/joc.1276 DOI: 10.1002/joc.1276]</ref>
 
−
|Delta||T(min, max,</br>avg), Pr||NetCDF||grid: 30 arc sec to</br>10 arc min||month
 
−
|-
 
−
| Bias Corrected / Spatial</br>Disaggregation (BCSD)<ref name="Wood2002">Wood, A.W., Maurer, E.P., Kumar, A. and Lettenmaier, D.P., 2002. Long‐range experimental hydrologic forecasting for the eastern United States. Journal of Geophysical Research: Atmospheres, 107(D20), 4429, pp. ACL6 1-15. [https://doi.org/10.1029/2001JD000659 DOI:10.1029/2001JD000659]&nbsp;&nbsp; Free access article available from: [https://agupubs.onlinelibrary.wiley.com/doi/abs/10.1029/2001JD000659 American Geophysical Union]&nbsp;&nbsp; [[Media: Wood2002.pdf | Report.pdf ]]</ref>
 
−
|Empirical Quantile</br>Mapping||Runoff,</br>Streamflow||NetCDF||grid: 7.5 arc min||day
 
−
|-
 
−
| [https://cida.usgs.gov/thredds/catalog.html?dataset=dcp Asynchronous Regional Regression</br>Model (ARRM v.1)]<ref name="Stoner2013">Stoner, A.M., Hayhoe, K., Yang, X., and Wuebbles, D.J., 2013. An Asynchronous Regional Regression Model for Statistical Downscaling of Daily Climate Variables. International Journal of Climatology, 33(11), pp. 2473-2494. [https://doi.org/10.1002/joc.3603 DOI:10.1002/joc.3603]</ref>
 
−
|Parameterized</br>Quantile Mapping||T(min, max), Pr||NetCDF||stations plus</br>grid: 7.5 arc min||day
 
−
|-
 
−
| [https://sdsm.org.uk/ Statistical Downscaling Model (SDSM)]<ref name="Wilby2013">Wilby, R.L., and Dawson, C.W., 2013. The Statistical DownScaling Model: insights from one decade of application. International Journal of Climatology, 33(7), pp. 1707-1719. [https://doi.org/10.1002/joc.3544 DOI: 10.1002/joc.3544]</ref>
 
−
|Weather Generator||T(min, max), Pr||PC Code||stations||day
 
−
|-
 
−
| [https://climate.northwestknowledge.net/MACA/ Multivariate Adaptive</br>Constructed Analogs (MACA)]<ref name="Hidalgo2008">Hidalgo, H.G., Dettinger, M.D. and Cayan, D.R., 2008. Downscaling with Constructed Analogues: Daily Precipitation and Temperature Fields Over the United States. California Energy Commission PIER Final Project, Report CEC-500-2007-123. [[Media: Hidalgo2008.PDF | Report.pdf]]</ref>
 
−
|Constructed Analogues||10 Variables||NetCDF||grid: 2.5 arc min||day
 
−
|-
 
−
| [http://loca.ucsd.edu/ Localized Constructed Analogs (LOCA)]<ref name="Pierce2013">Pierce, D.W., Cayan, D.R. and Thrasher, B.L., 2014. Statistical Downscaling Using Localized Constructed Analogs (LOCA). Journal of Hydrometeorology, 15(6), pp. 2558-2585. [https://doi.org/10.1175/JHM-D-14-0082.1 DOI: 10.1175/JHM-D-14-0082.1]&nbsp;&nbsp; Free access article available from: [https://journals.ametsoc.org/view/journals/hydr/15/6/jhm-d-14-0082_1.xml American Meteorological Society].&nbsp;&nbsp; [[Media: Pierce2014.pdf | Report.pdf]]</ref>
 
−
|Constructed Analogues||T(min, max), Pr||NetCDF||grid: 3.75 arc min||day
 
−
|-
 
−
| [https://www.nccs.nasa.gov/services/data-collections/land-based-products/nex-dcp30 NASA Earth Exchange Downscaled</br>Climate Projections (NEX-DCP30)]<ref name="Wood2002"/>
 
−
|Bias Correction /</br>Spatial Disaggregation||T(min, max), Pr||NetCDF||grid: 30 arc sec||month
 
−
|-
 
−
| colspan="6" style="text-align: left; background-color:white;" |'''Dynamical Downscaled Datasets'''
 
−
|-
 
−
| [http://www.narccap.ucar.edu/index.html North American Regional Climate</br>Change Assessment Program (NARCCAP)]<ref name="Mearns2009">Mearns, L.O., Gutowski, W., Jones, R., Leung, R., McGinnis, S., Nunes, A. and Qian, Y., 2009. A Regional Climate Change Assessment Program for North America. Eos, Transactions, American Geophysical Union, 90(36), p.311.  [https://doi.org/10.1029/2009EO360002 DOI: 10.1029/2009EO360002]&nbsp;&nbsp; Free access article from: [https://agupubs.onlinelibrary.wiley.com/doi/abs/10.1029/2009EO360002 American Geophysical Union]&nbsp;&nbsp; [[Media: Mearns2009.pdf | Report.pdf]]</ref>
 
−
|Multiple Models||49 Variables||NetCDF||grid: 30 arc min||3 hours
 
−
|-
 
−
| [https://cordex.org/about/ Coordinated Regional Climate</br>Downscaling Experiment (CORDEX)]<ref name="Giorgi2009">Giorgi, F., Jones, C., and Asrar, G.R., 2009. Addressing climate information needs at the regional level: the CORDEX framework. World Meteorological Organization (WMO) Bulletin, 58(3), pp. 175-183. Free access article from: [https://public.wmo.int/en/bulletin/addressing-climate-information-needs-regional-level-cordex-framework World Meteorological Organization]&nbsp;&nbsp; [[Media: Giorgi2009.pdf | Report.pdf]]</ref>
 
−
|Multiple Models||66 Variables||NetCDF||grid: 30 arc min||3 hours
 
−
|-
 
−
| [https://esrl.noaa.gov/gsd/wrfportal/ Strategic Environmental Research and</br>Development Program (SERDP)]<ref name="Wang2015">Wang, J., and Kotamarthi, V.R., 2015. High‐resolution dynamically downscaled projections of precipitation in the mid and late 21st century over North America. Earth's Future, 3(7), pp. 268-288.  [https://doi.org/10.1002/2015EF000304 DOI: 10.1002/2015EF000304]&nbsp;&nbsp; Free access article from: [https://agupubs.onlinelibrary.wiley.com/doi/full/10.1002/2015EF000304 American Geophysical Union]&nbsp;&nbsp; [[Media: Wang2015.pdf | Report.pdf]]</ref>
 
−
|Weather Research and</br>Forecasting (WRF v3.3)||80+ Variables||NetCDF||grid: 6.5 arc min||3 hours
 
−
|}
 
−
A&nbsp;primary&nbsp;cause&nbsp;of&nbsp;uncertainty in climate change projections, especially beyond 30 years into the future, is the uncertainty in the greenhouse gas (GHG) emission scenarios used to make climate model projections. The best method of accounting for this type of uncertainty is to apply a climate change model to multiple GHG emission scenarios (see also: [[Wikipedia: Representative Concentration Pathway]]).  
 
  
−
The&nbsp;uncertainties&nbsp;in&nbsp;climate&nbsp;projections over shorter timescales, less than 30 years out, are dominated by something known as “internal variability” in the models. Different approaches are used to address the uncertainty from internal variability<ref name="Kotamarthi2021"/>. A third type of uncertainty in climate modeling, known as scientific uncertainty, comes from our inability to numerically solve every aspect of the complex earth system. We expect this scientific uncertainty to decrease as we understand more of the earth system and improve its representation in our numerical models. As discussed in [[Climate Change Primer]], numerical experiments based on global climate models are designed to address these uncertainties in various ways. Downscaling methods evaluate this uncertainty by using several independent regional climate models to generate future projections, with the expectation that each of these models will capture some aspects of the physics better than the others, and that by using several different models, we can estimate the range of this uncertainty. Thus, the commonly accepted methods for accounting for uncertainty in climate model projections are either using projections from one model for several emission scenarios, or applying multiple models to project a single scenario.  
+
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>.
  
−
A comparison of the currently available methods and their characteristics is provided in Table 2 (adapted from Kotamarthi et al., 2016<ref name="Kotamarthi2016"/>).  The table lists the various methodologies and models used for producing downscaled data, and the climate variables that these methods produce.  These datasets are mostly available for download from the data servers and websites listed in the table and in a few cases by contacting the respective source organizations.
+
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"/>.
  
−
The most popular and widely used format for atmospheric and climate science is known as [[Wikipedia:NetCDF | NetCDF]], which stands for Network Common Data Form. NetCDF is a self-describing data format that saves data in a binary format. The format is self-describing in that a metadata listing is part of every file that describes all the data attributes, such as dimensions, units and data size and in principal should not need additional information to extract the required data for analysis with the right software. However, specially built software for reading and extracting data from these binary files is necessary for making visualizations and further analysis. Software packages for reading and writing NetCDF datasets and for generating visualizations from these datasets are widely available and obtained free of cost ([https://www.unidata.ucar.edu/software/netcdf/docs/ NetCDF-tools]). Popular geospatial analysis tools such as ARC-GIS, statistical packages such as ‘R’ and programming languages such as Fortran, C++, and Python have built in libraries that can be used to directly read NetCDF files for visualization and analysis.  
+
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"/>.
−
<br clear="left" />
+
 
 +
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"/>.
 +
 
 +
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]
−
[https://serdp-estcp.org/Program-Areas/Resource-Conservation-and-Resiliency/Infrastructure-Resiliency/Vulnerability-and-Impact-Assessment/RC-2242/(language)/eng-US Climate Change Impacts to Department of Defense Installations, SERDP Project RC-2242]
 

Latest revision as of 18:26, 15 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

File:FullerFig1.png
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

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].

Toxicological Effects of PFAS

The characterization of toxicological effects in human health risk assessments is based on toxicological studies of mammalian exposures to per- and polyfluoroalkyl substances (PFAS), primarily studies involving perfluorooctanesulfonic acid (PFOS) and perfluorooctanoic acid (PFOA). The most sensitive noncancer adverse effects involve the liver and kidney, immune system, and various developmental and reproductive endpoints[7]. A select number of PFAS have been evaluated for carcinogenicity, primarily using epidemiological data. Only PFOS and PFOA (and their derivatives) have sufficient data for USEPA to characterize as Likely to Be Carcinogenic to Humans via the oral route of exposure. Epidemiological studies provided evidence of bladder, prostate, liver, kidney, and breast cancers in humans related to PFOS exposure, as well as kidney and testicular cancer in humans and limited evidence of breast cancer related to PFOA exposure[7][8][9].

USEPA’s Integrated Risk Management System (IRIS) Program is developing Toxicological Reviews to improve understanding of the toxicity of several additional PFAS (i.e., not solely PFOA and PFOS). Toxicological Reviews provide an overview of cancer and noncancer health effects based on current literature and, where data are sufficient, derive human health toxicity criteria (i.e., human health oral reference doses and cancer slope factors) that form the basis for risk-based decision making. For risk assessors, these documents provide USEPA reference doses and cancer slope factors that can be used with exposure information and other considerations to assess human health risk. Final Toxicological Reviews have been completed for the following PFAS:

  • Perfluorooctanesulfonic acid (PFOS)
  • Perfluorooctanoic acid (PFOA)
  • Perfluorobutanoic acid (PFBA)
  • Perfluorohexanoic acid (PFHxA)
  • Perfluorobutane sulfonic acid (PFBS)
  • Perfluoropropionic acid (PFPrA)
  • Perfluorohexane sulfonic acid (PFHxS)
  • Lithium bis[(trifluoromethyl)sulfonyl]azanide (HQ-115)
  • Hexafluoropropylene oxide dimer acid (HFPO DA) and its Ammonium Salt

Toxicity assessments are ongoing for the following PFAS:

  • Perfluorononanoic acid (PFNA)
  • Perfluorodecanoic acid (PFDA)

It is important to note human health toxicity criteria for inhalation of PFAS are not included in the Final Toxicological Reviews and are not currently available. In addition to IRIS, state agencies have developed peer-reviewed provisional toxicity values that have been incorporated into USEPA’s RSLs, which are updated biannually. These values have not been reviewed by or incorporated into IRIS.

With respect to ecological toxicity, effects on reproduction, growth, and development of avian and mammalian wildlife have been documented in controlled laboratory studies of exposures of standard toxicological test species (e.g., mice, quail) to PFAS. Many of these studies have been reviewed[10][11][12][13] to derive ecological Toxicity Reference Values (TRVs). TRVs can be used alongside exposure information and other considerations to assess ecological risk. Avian and mammalian wildlife receptors are generally expected to have the highest risks due to PFAS exposure. Direct toxicity to aquatic life, such as fish and invertebrates, from exposure to sediment and surface water also occurs, though concentrations in water 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 less sensitive to PFAS when compared to terrestrial wildlife, with risk-based PFAS concentrations in soil being much higher than those associated with potential effects to terrestrial wildlife[13].

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[7]:

  • 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[14]: 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[14][13][10]. USEPA has developed water quality criteria for aquatic life[15][16][17] for PFOA and PFOS. Following extensive reviews of the peer-reviewed literature, Zodrow et al.[13] used the USEPA Great Lakes Initiative methodology[18] 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[19]. 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[10][13], and the Canadian Council of Ministers of Environment developed several draft thresholds protective of direct toxicity of PFOS in soil[20].

Wildlife screening levels for abiotic media are back-calculated from food web models developed for representative receptors. Both Zodrow et al.[13] and Grippo et al.[19] include the development of risk-based screening levels for wildlife. The Michigan Department of Community Health[21] 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.[13]. 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[22].

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[23][24]. 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[14], 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[25][26][27][28][29][30][31][32][33][34]. In addition to fish, terrestrial wildlife can accumulate contaminants from impacted sites, resulting in potential exposures to consumers of wild game[35]. 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[36]. Multiple studies have found PFAS can be taken up by plants from soil porewater[37][38][39], and livestock can accumulate PFAS from drinking water and/or feed[40]. 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[8][9] 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[41]. 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[14].

General and site-specific human health exposure pathways and risk assessment methods as outlined by USEPA[42][43] 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[44][10][13]. 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[13].

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[13][45].

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[46][44].

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[43].

Recently, Conder et al.[10], Gobas et al.[11], and Zodrow et al.[13] 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[10][11][13][47][12]. 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[44][10][11][12][13].

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[43] 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
  2. ^ 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. doi: 10.1016/j.jhydrol.2005.05.021
  3. ^ 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. doi: 10.1016/j.watres.2015.10.019
  4. ^ 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. doi: 10.2166/wst.2006.617
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See Also

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

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