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===Biochar===
 
===Biochar===
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Recent reports have highlighted additional materials that, either alone, or in combination with electron donors such as peat moss and soybean oil, may further enhance the sorption and degradation of surface runoff contaminants, including both legacy energetics and [[Wikipedia: Insensitive_munition#Insensitive_high_explosives | insensitive high explosives (IHE)]].  For instance, [[Wikipedia: Biochar | biochar]], a type of black carbon, has been shown to not only sorb a wide range of organic and inorganic contaminants including MCs<ref>Ahmad, M., Rajapaksha, A.U., Lim, J.E., Zhang, M., Bolan, N., Mohan, D., Vithanage, M., Lee, S.S., Ok, Y.S., 2014. Biochar as a sorbent for contaminant management in soil and water: A review. Chemosphere, 99, pp. 19-33. [https://doi.org/10.1016/j.chemosphere.2013.10.071 doi: 10.1016/j.chemosphere.2013.10.071]</ref><ref>Mohan, D., Sarswat, A., Ok, Y.S., Pittman, C.U., 2014. Organic and inorganic contaminants removal from water with biochar, a renewable, low cost and sustainable adsorbent – A critical review. Bioresource Technology, 160, pp. 191-202. [https://doi.org/10.1016/j.biortech.2014.01.120 doi: 10.1016/j.biortech.2014.01.120]</ref><ref>Oh, S.-Y., Seo, Y.-D., Jeong, T.-Y., Kim, S.-D., 2018. Sorption of Nitro Explosives to Polymer/Biomass-Derived Biochar. Journal of Environmental Quality, 47(2), pp. 353-360. [https://doi.org/10.2134/jeq2017.09.0357 doi: 10.2134/jeq2017.09.0357]</ref><ref>Xie, T., Reddy, K.R., Wang, C., Yargicoglu, E., Spokas, K., 2015. Characteristics and Applications of Biochar for Environmental Remediation: A Review. Critical Reviews in Environmental Science and Technology, 45(9), pp. 939-969. [https://doi.org/10.1080/10643389.2014.924180 doi: 10.1080/10643389.2014.924180]</ref>, but also to facilitate their degradation<ref>Oh, S.-Y., Cha, D.K., Kim, B.-J., Chiu, P.C., 2002. Effect of adsorption to elemental iron on the transformation of 2,4,6-trinitrotoluene and hexahydro-1,3,5-trinitro-1,3,5-triazine in solution. Environmental Toxicology and Chemistry, 21(7), pp. 1384-1389. [https://doi.org/10.1002/etc.5620210708 doi: 10.1002/etc.5620210708]</ref><ref>Ye, J., Chiu, P.C., 2006. Transport of Atomic Hydrogen through Graphite and its Reaction with Azoaromatic Compounds. Environmental Science and Technology, 40(12), pp. 3959-3964. [https://doi.org/10.1021/es060038x doi: 10.1021/es060038x]</ref><ref name="OhChiu2009">Oh, S.-Y., Chiu, P.C., 2009. Graphite- and Soot-Mediated Reduction of 2,4-Dinitrotoluene and Hexahydro-1,3,5-trinitro-1,3,5-triazine. Environmental Science and Technology, 43(18), pp. 6983-6988. [https://doi.org/10.1021/es901433m doi: 10.1021/es901433m]</ref><ref name="OhEtAl2013">Oh, S.-Y., Son, J.-G., Chiu, P.C., 2013. Biochar-mediated reductive transformation of nitro herbicides and explosives. Environmental Toxicology and Chemistry, 32(3), pp. 501-508. [https://doi.org/10.1002/etc.2087 doi: 10.1002/etc.2087]&nbsp;&nbsp; [[Media: OhEtAl2013.pdf | Open Access Article.pdf]]</ref><ref name="XuEtAl2010">Xu, W., Dana, K.E., Mitch, W.A., 2010. Black Carbon-Mediated Destruction of Nitroglycerin and RDX by Hydrogen Sulfide. Environmental Science and Technology, 44(16), pp. 6409-6415. [https://doi.org/10.1021/es101307n doi: 10.1021/es101307n]</ref><ref>Xu, W., Pignatello, J.J., Mitch, W.A., 2013. Role of Black Carbon Electrical Conductivity in Mediating Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) Transformation on Carbon Surfaces by Sulfides. Environmental Science and Technology, 47(13), pp. 7129-7136. [https://doi.org/10.1021/es4012367 doi: 10.1021/es4012367]</ref>. Depending on the source biomass and [[Wikipedia: Pyrolysis| pyrolysis]] conditions, biochar can possess a high [[Wikipedia: Specific surface area | specific surface area]] (on the order of several hundred m<small><sup>2</sup></small>/g)<ref>Zhang, J., You, C., 2013. Water Holding Capacity and Absorption Properties of Wood Chars. Energy and Fuels, 27(5), pp. 2643-2648. [https://doi.org/10.1021/ef4000769 doi: 10.1021/ef4000769]</ref><ref>Gray, M., Johnson, M.G., Dragila, M.I., Kleber, M., 2014. Water uptake in biochars: The roles of porosity and hydrophobicity. Biomass and Bioenergy, 61, pp. 196-205. [https://doi.org/10.1016/j.biombioe.2013.12.010 doi: 10.1016/j.biombioe.2013.12.010]</ref> and hence a high sorption capacity.  Biochar and other black carbon also exhibit especially high affinity for [[Wikipedia: Nitro compound | nitroaromatic compounds (NACs)]] including TNT and 2,4-dinitrotoluene (DNT)<ref>Sander, M., Pignatello, J.J., 2005. Characterization of Charcoal Adsorption Sites for Aromatic Compounds:  Insights Drawn from Single-Solute and Bi-Solute Competitive Experiments. Environmental Science and Technology, 39(6), pp. 1606-1615. [https://doi.org/10.1021/es049135l doi: 10.1021/es049135l]</ref><ref name="ZhuEtAl2005">Zhu, D., Kwon, S., Pignatello, J.J., 2005. Adsorption of Single-Ring Organic Compounds to Wood Charcoals Prepared Under Different Thermochemical Conditions. Environmental Science and Technology 39(11), pp. 3990-3998. [https://doi.org/10.1021/es050129e doi: 10.1021/es050129e]</ref><ref name="ZhuPignatello2005">Zhu, D., Pignatello, J.J., 2005. Characterization of Aromatic Compound Sorptive Interactions with Black Carbon (Charcoal) Assisted by Graphite as a Model. Environmental Science and Technology, 39(7), pp. 2033-2041. [https://doi.org/10.1021/es0491376 doi: 10.1021/es0491376]</ref>. This is due to the strong [[Wikipedia: Pi-interaction | ''&pi;-&pi;'' electron donor-acceptor interactions]] between electron-rich graphitic domains in black carbon and the electron-deficient aromatic ring of the NAC
+
Recent reports have highlighted additional materials that, either alone, or in combination with electron donors such as peat moss and soybean oil, may further enhance the sorption and degradation of surface runoff contaminants, including both legacy energetics and [[Wikipedia: Insensitive_munition#Insensitive_high_explosives | insensitive high explosives (IHE)]].  For instance, [[Wikipedia: Biochar | biochar]], a type of black carbon, has been shown to not only sorb a wide range of organic and inorganic contaminants including MCs<ref>Ahmad, M., Rajapaksha, A.U., Lim, J.E., Zhang, M., Bolan, N., Mohan, D., Vithanage, M., Lee, S.S., Ok, Y.S., 2014. Biochar as a sorbent for contaminant management in soil and water: A review. Chemosphere, 99, pp. 19-33. [https://doi.org/10.1016/j.chemosphere.2013.10.071 doi: 10.1016/j.chemosphere.2013.10.071]</ref><ref>Mohan, D., Sarswat, A., Ok, Y.S., Pittman, C.U., 2014. Organic and inorganic contaminants removal from water with biochar, a renewable, low cost and sustainable adsorbent – A critical review. Bioresource Technology, 160, pp. 191-202. [https://doi.org/10.1016/j.biortech.2014.01.120 doi: 10.1016/j.biortech.2014.01.120]</ref><ref>Oh, S.-Y., Seo, Y.-D., Jeong, T.-Y., Kim, S.-D., 2018. Sorption of Nitro Explosives to Polymer/Biomass-Derived Biochar. Journal of Environmental Quality, 47(2), pp. 353-360. [https://doi.org/10.2134/jeq2017.09.0357 doi: 10.2134/jeq2017.09.0357]</ref><ref>Xie, T., Reddy, K.R., Wang, C., Yargicoglu, E., Spokas, K., 2015. Characteristics and Applications of Biochar for Environmental Remediation: A Review. Critical Reviews in Environmental Science and Technology, 45(9), pp. 939-969. [https://doi.org/10.1080/10643389.2014.924180 doi: 10.1080/10643389.2014.924180]</ref>, but also to facilitate their degradation<ref>Oh, S.-Y., Cha, D.K., Kim, B.-J., Chiu, P.C., 2002. Effect of adsorption to elemental iron on the transformation of 2,4,6-trinitrotoluene and hexahydro-1,3,5-trinitro-1,3,5-triazine in solution. Environmental Toxicology and Chemistry, 21(7), pp. 1384-1389. [https://doi.org/10.1002/etc.5620210708 doi: 10.1002/etc.5620210708]</ref><ref>Ye, J., Chiu, P.C., 2006. Transport of Atomic Hydrogen through Graphite and its Reaction with Azoaromatic Compounds. Environmental Science and Technology, 40(12), pp. 3959-3964. [https://doi.org/10.1021/es060038x doi: 10.1021/es060038x]</ref><ref name="OhChiu2009">Oh, S.-Y., Chiu, P.C., 2009. Graphite- and Soot-Mediated Reduction of 2,4-Dinitrotoluene and Hexahydro-1,3,5-trinitro-1,3,5-triazine. Environmental Science and Technology, 43(18), pp. 6983-6988. [https://doi.org/10.1021/es901433m doi: 10.1021/es901433m]</ref><ref name="OhEtAl2013">Oh, S.-Y., Son, J.-G., Chiu, P.C., 2013. Biochar-mediated reductive transformation of nitro herbicides and explosives. Environmental Toxicology and Chemistry, 32(3), pp. 501-508. [https://doi.org/10.1002/etc.2087 doi: 10.1002/etc.2087]&nbsp;&nbsp; [[Media: OhEtAl2013.pdf | Open Access Article.pdf]]</ref><ref name="XuEtAl2010">Xu, W., Dana, K.E., Mitch, W.A., 2010. Black Carbon-Mediated Destruction of Nitroglycerin and RDX by Hydrogen Sulfide. Environmental Science and Technology, 44(16), pp. 6409-6415. [https://doi.org/10.1021/es101307n doi: 10.1021/es101307n]</ref><ref>Xu, W., Pignatello, J.J., Mitch, W.A., 2013. Role of Black Carbon Electrical Conductivity in Mediating Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) Transformation on Carbon Surfaces by Sulfides. Environmental Science and Technology, 47(13), pp. 7129-7136. [https://doi.org/10.1021/es4012367 doi: 10.1021/es4012367]</ref>. Depending on the source biomass and [[Wikipedia: Pyrolysis| pyrolysis]] conditions, biochar can possess a high [[Wikipedia: Specific surface area | specific surface area]] (on the order of several hundred m<small><sup>2</sup></small>/g)<ref>Zhang, J., You, C., 2013. Water Holding Capacity and Absorption Properties of Wood Chars. Energy and Fuels, 27(5), pp. 2643-2648. [https://doi.org/10.1021/ef4000769 doi: 10.1021/ef4000769]</ref><ref>Gray, M., Johnson, M.G., Dragila, M.I., Kleber, M., 2014. Water uptake in biochars: The roles of porosity and hydrophobicity. Biomass and Bioenergy, 61, pp. 196-205. [https://doi.org/10.1016/j.biombioe.2013.12.010 doi: 10.1016/j.biombioe.2013.12.010]</ref> and hence a high sorption capacity.  Biochar and other black carbon also exhibit especially high affinity for [[Wikipedia: Nitro compound | nitroaromatic compounds (NACs)]] including TNT and 2,4-dinitrotoluene (DNT)<ref>Sander, M., Pignatello, J.J., 2005. Characterization of Charcoal Adsorption Sites for Aromatic Compounds:  Insights Drawn from Single-Solute and Bi-Solute Competitive Experiments. Environmental Science and Technology, 39(6), pp. 1606-1615. [https://doi.org/10.1021/es049135l doi: 10.1021/es049135l]</ref><ref name="ZhuEtAl2005">Zhu, D., Kwon, S., Pignatello, J.J., 2005. Adsorption of Single-Ring Organic Compounds to Wood Charcoals Prepared Under Different Thermochemical Conditions. Environmental Science and Technology 39(11), pp. 3990-3998. [https://doi.org/10.1021/es050129e doi: 10.1021/es050129e]</ref><ref name="ZhuPignatello2005">Zhu, D., Pignatello, J.J., 2005. Characterization of Aromatic Compound Sorptive Interactions with Black Carbon (Charcoal) Assisted by Graphite as a Model. Environmental Science and Technology, 39(7), pp. 2033-2041. [https://doi.org/10.1021/es0491376 doi: 10.1021/es0491376]</ref>. This is due to the strong [[Wikipedia: Pi-interaction | ''&pi;-&pi;'' electron donor-acceptor interactions]] between electron-rich graphitic domains in black carbon and the electron-deficient aromatic ring of the NAC<ref name="ZhuEtAl2005"/><ref name="ZhuPignatello2005"/>. These characteristics make biochar a potentially effective, low cost, and sustainable sorbent for removing MC and other contaminants from surface runoff and retaining them for subsequent degradation ''in situ''.
 +
 
 +
Furthermore, black carbon such as biochar can promote abiotic and microbial transformation reactions by facilitating electron transfer.  That is, biochar is not merely a passive sorbent for contaminants, but also a redox mediator for their degradation.  Biochar can promote contaminant degradation through two different mechanisms: electron conduction and electron storage<ref>Sun, T., Levin, B.D.A., Guzman, J.J.L., Enders, A., Muller, D.A., Angenent, L.T., Lehmann, J., 2017. Rapid electron transfer by the carbon matrix in natural pyrogenic carbon. Nature Communications, 8, Article 14873. [https://doi.org/10.1038/ncomms14873 doi: 10.1038/ncomms14873]&nbsp;&nbsp; [[Media: SunEtAl2017.pdf | Open Access Article.pdf]]</ref>.
 +
 
 +
First, the microscopic graphitic regions in biochar can adsorb contaminants like NACs strongly, as noted above, and also conduct reducing equivalents such as electrons and atomic hydrogen to the sorbed contaminants, thus promoting their reductive degradation.  This catalytic process has been demonstrated for TNT, DNT, RDX, HMX, and [[Wikipedia: Nitroglycerin | nitroglycerin]]<ref>Oh, S.-Y., Cha, D.K., Chiu, P.C., 2002. Graphite-Mediated Reduction of 2,4-Dinitrotoluene with Elemental Iron. Environmental Science and Technology, 36(10), pp. 2178-2184. [https://doi.org/10.1021/es011474g doi: 10.1021/es011474g]</ref><ref>Oh, S.-Y., Cha, D.K., Kim, B.J., Chiu, P.C., 2004. Reduction of Nitroglycerin with Elemental Iron:  Pathway, Kinetics, and Mechanisms. Environmental Science and Technology, 38(13), pp. 3723-3730. [https://doi.org/10.1021/es0354667 doi: 10.1021/es0354667]</ref><ref>Oh, S.-Y., Cha, D.K., Kim, B.J., Chiu, P.C., 2005. Reductive transformation of hexahydro-1,3,5-trinitro-1,3,5-triazine, octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine, and methylenedinitramine with elemental iron. Environmental Toxicology and Chemistry, 24(11), pp. 2812-2819. [https://doi.org/10.1897/04-662R.1 doi: 10.1897/04-662R.1]</ref><ref name="OhChiu2009"/><ref name="XuEtAl2010"/> and is expected to occur also for IHE including DNAN and NTO.
 +
 
 +
Second, biochar contains in its structure abundant redox-facile functional groups such as [[Wikipedia: Quinone | quinones]] and [[Wikipedia: Hydroquinone | hydroquinones]], which are known to accept and donate electrons reversibly.  Depending on the biomass and pyrolysis temperature, certain biochar can possess a rechargeable electron storage capacity (i.e., reversible electron accepting and donating capacity) on the order of several millimoles e<small><sup>–</sup></small>/g<ref>Klüpfel, L., Keiluweit, M., Kleber, M., Sander, M., 2014. Redox Properties of Plant Biomass-Derived Black Carbon (Biochar). Environmental Science and Technology, 48(10), pp. 5601-5611. [https://doi.org/10.1021/es500906d doi: 10.1021/es500906d]</ref><ref>Prévoteau, A., Ronsse, F., Cid, I., Boeckx, P., Rabaey, K., 2016. The electron donating capacity of biochar is dramatically underestimated. Scientific Reports, 6, Article 32870. [https://doi.org/10.1038/srep32870 doi: 10.1038/srep32870]&nbsp;&nbsp; [[Media: PrevoteauEtAl2016.pdf | Open Access Article.pdf]]</ref><ref>Xin, D., Xian, M., Chiu, P.C., 2018. Chemical methods for determining the electron storage capacity of black carbon. MethodsX, 5, pp. 1515-1520. [https://doi.org/10.1016/j.mex.2018.11.007 doi: 10.1016/j.mex.2018.11.007]&nbsp;&nbsp; [[Media: XinEtAl2018.pdf | Open Access Article.pdf]]</ref>. This means that when "charged", biochar can provide electrons for either abiotic or biotic degradation of reducible compounds such as MC. The abiotic reduction of DNT and RDX mediated by biochar has been demonstrated<ref name="OhEtAl2013"/> and similar reactions are expected to occur for DNAN and NTO as well. Recent studies have shown that the electron storage capacity of biochar is also accessible to microbes.  For example, soil bacteria such as [[Wikipedia: Geobacter | Geobacter]] and [[Wikipedia: Shewanella | Shewanella]] species can utilize oxidized (or "discharged") biochar as an electron acceptor for the oxidation of organic substrates such as lactate and acetate<ref>Kappler, A., Wuestner, M.L., Ruecker, A., Harter, J., Halama, M., Behrens, S., 2014. Biochar as an Electron Shuttle between Bacteria and Fe(III) Minerals. Environmental Science and Technology Letters, 1(8), pp. 339-344. [https://doi.org/10.1021/ez5002209 doi: 10.1021/ez5002209]</ref><ref name="SaquingEtAl2016">Saquing, J.M., Yu, Y.-H., Chiu, P.C., 2016. Wood-Derived Black Carbon (Biochar) as a Microbial Electron Donor and Acceptor. Environmental Science and Technology Letters, 3(2), pp. 62-66. [https://doi.org/10.1021/acs.estlett.5b00354 doi: 10.1021/acs.estlett.5b00354]</ref> and reduced (or "charged") biochar as an electron donor for the reduction of nitrate<ref name="SaquingEtAl2016"/>. This is significant because, through microbial access of stored electrons in biochar, contaminants that do not sorb strongly to biochar can still be degraded.
 +
 
 +
Similar to nitrate, perchlorate and other relatively water-soluble energetic compounds (e.g., NTO and NQ) may also be similarly transformed using reduced biochar as an electron donor.  Unlike other electron donors, biochar can be recharged through biodegradation of organic substrates<ref name="SaquingEtAl2016"/> and thus can serve as a long-lasting sorbent and electron repository in soil.  Similar to peat moss, the high porosity and surface area of biochar not only facilitate contaminant sorption but also create anaerobic reducing microenvironments in its inner pores, where reductive degradation of energetic compounds can take place. 
 +
 
 +
===Other Sorbents===
 +
Chitin and unmodified cellulose were predicted by [[Wikipedia: Density functional theory | Density Functional Theory]] methods to be favorable for absorption of NTO and NQ, as well as the legacy explosives
  
 
==PFAS Screening Levels for Human Health and Ecological Risk Assessments==
 
==PFAS Screening Levels for Human Health and Ecological Risk Assessments==

Latest revision as of 18:46, 21 October 2025

Remediation of Stormwater Runoff Contaminated by Munition Constituents

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

Related Article(s):


Contributor: Mark E. Fuller

Key Resource(s):

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

Background

Surface Runoff Characteristics and Treatment Approaches

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

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

Surface Runoff on Ranges

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

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

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

Range Runoff Treatment Technology Components

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

Peat

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

Soybean Oil

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

Biochar

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

Furthermore, black carbon such as biochar can promote abiotic and microbial transformation reactions by facilitating electron transfer. That is, biochar is not merely a passive sorbent for contaminants, but also a redox mediator for their degradation. Biochar can promote contaminant degradation through two different mechanisms: electron conduction and electron storage[33].

First, the microscopic graphitic regions in biochar can adsorb contaminants like NACs strongly, as noted above, and also conduct reducing equivalents such as electrons and atomic hydrogen to the sorbed contaminants, thus promoting their reductive degradation. This catalytic process has been demonstrated for TNT, DNT, RDX, HMX, and nitroglycerin[34][35][36][24][26] and is expected to occur also for IHE including DNAN and NTO.

Second, biochar contains in its structure abundant redox-facile functional groups such as quinones and hydroquinones, which are known to accept and donate electrons reversibly. Depending on the biomass and pyrolysis temperature, certain biochar can possess a rechargeable electron storage capacity (i.e., reversible electron accepting and donating capacity) on the order of several millimoles e–/g[37][38][39]. This means that when "charged", biochar can provide electrons for either abiotic or biotic degradation of reducible compounds such as MC. The abiotic reduction of DNT and RDX mediated by biochar has been demonstrated[25] and similar reactions are expected to occur for DNAN and NTO as well. Recent studies have shown that the electron storage capacity of biochar is also accessible to microbes. For example, soil bacteria such as Geobacter and Shewanella species can utilize oxidized (or "discharged") biochar as an electron acceptor for the oxidation of organic substrates such as lactate and acetate[40][41] and reduced (or "charged") biochar as an electron donor for the reduction of nitrate[41]. This is significant because, through microbial access of stored electrons in biochar, contaminants that do not sorb strongly to biochar can still be degraded.

Similar to nitrate, perchlorate and other relatively water-soluble energetic compounds (e.g., NTO and NQ) may also be similarly transformed using reduced biochar as an electron donor. Unlike other electron donors, biochar can be recharged through biodegradation of organic substrates[41] and thus can serve as a long-lasting sorbent and electron repository in soil. Similar to peat moss, the high porosity and surface area of biochar not only facilitate contaminant sorption but also create anaerobic reducing microenvironments in its inner pores, where reductive degradation of energetic compounds can take place.

Other Sorbents

Chitin and unmodified cellulose were predicted by Density Functional Theory methods to be favorable for absorption of NTO and NQ, as well as the legacy explosives

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[42]:

  • 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[43]: 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[43][44][45]. USEPA has developed water quality criteria for aquatic life[46][47][48] for PFOA and PFOS. Following extensive reviews of the peer-reviewed literature, Zodrow et al.[44] used the USEPA Great Lakes Initiative methodology[49] 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[50]. 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[45][44], and the Canadian Council of Ministers of Environment developed several draft thresholds protective of direct toxicity of PFOS in soil[51].

Wildlife screening levels for abiotic media are back-calculated from food web models developed for representative receptors. Both Zodrow et al.[44] and Grippo et al.[50] include the development of risk-based screening levels for wildlife. The Michigan Department of Community Health[52] 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.[44]. 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[53].

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[54][55]. 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[43], 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[56][57][58][59][60][61][62][63][64][65]. In addition to fish, terrestrial wildlife can accumulate contaminants from impacted sites, resulting in potential exposures to consumers of wild game[66]. 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[67]. Multiple studies have found PFAS can be taken up by plants from soil porewater[68][69][70], and livestock can accumulate PFAS from drinking water and/or feed[71]. 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[72][73] 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[74]. 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[43].

General and site-specific human health exposure pathways and risk assessment methods as outlined by USEPA[75][76] 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[77][45][44]. 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[44].

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[44][78].

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[79][77].

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[76].

Recently, Conder et al.[45], Gobas et al.[80], and Zodrow et al.[44] 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[45][80][44][81][82]. 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[77][45][80][82][44].

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[76] 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.

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See Also

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

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