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==Passive Sampling of Sediments==
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==PFAS Treatment by Anion Exchange==  
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"Passive sampling" refers to a group of methods used to quantify the availability of organic contaminants to move between different media and/or to react in environmental systems such as indoor air, lake waters, or contaminated sediment beds. To do this, the passive sampling material is deployed in the environmental system and allowed to absorb chemicals of interest via diffusive transfers from the surroundings.  Upon recovery of the passive sampler, the accumulated contaminants are measured, and the concentrations in the sampler are interpreted to infer the chemical concentrations in specific surrounding media like porewater in a sediment bed.  Such data are then useful inputs for site assessments such as those seeking to quantify fluxes from contaminated sediment beds to overlying waters or to evaluate the risk of significant uptake into benthic infauna and the larger food web.
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[[Wikipedia: Ion exchange | Anion exchange]] has emerged as one of the most effective and economical technologies for treatment of water contaminated by [[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) | per- and polyfluoroalkyl substances (PFAS)]]. Anion exchange resins (AERs) are polymer beads (0.5–1 mm diameter) incorporating cationic adsorption sites that attract anionic PFAS by a combination of electrostatic and hydrophobic mechanisms. Both regenerable and single-use resin treatment systems are being investigated, and results from pilot-scale studies show that AERs can treat much greater volumes of PFAS-contaminated water than comparable amounts of [[Wikipedia: Activated carbon | granular activated carbon (GAC)]] adsorbent media. Life cycle treatment costs and environmental impacts of anion exchange and other adsorbent technologies are highly dependent upon the treatment criteria selected by site managers to determine when media is exhausted and requires replacement or regeneration.
 
<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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* [[Contaminated Sediments - Introduction]]
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*[[Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS)]]  
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* [[In Situ Treatment of Contaminated Sediments with Activated Carbon]]
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*[[PFAS Sources]]
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* [[Passive Sampling of Munitions Constituents]]
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*[[PFAS Transport and Fate]]
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*[[PFAS Ex Situ Water Treatment]]
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*[[Supercritical Water Oxidation (SCWO)]]
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*[[PFAS Treatment by Electrical Discharge Plasma]]
  
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'''Contributor(s):''' [[Dr. Philip M. Gschwend]]
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'''Contributor(s):'''  
 +
*Dr. Timothy J. Strathmann
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*Dr. Anderson Ellis
 +
*Dr. Treavor H. Boyer
  
 
'''Key Resource(s):'''
 
'''Key Resource(s):'''
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* Validating the Use of Performance Reference Compounds in Passive Samplers to Assess Porewater Concentrations in Sediment Beds<ref name ="Apell2014">Apell, J.N. and Gschwend, P.M., 2014. Validating the Use of Performance Reference Compounds in Passive Samplers to Assess Porewater Concentrations in Sediment Beds.  Environmental Science and Technology, 48(17), pp. 10301-10307. [https://doi.org/10.1021/es502694g DOI: 10.1021/es502694g]</ref>
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*Anion Exchange Resin Removal of Per- and Polyfluoroalkyl Substances (PFAS) from Impacted Water: A Critical Review<ref name="BoyerEtAl2021a">Boyer, T.H., Fang, Y., Ellis, A., Dietz, R., Choi, Y.J., Schaefer, C.E., Higgins, C.P., Strathmann, T.J., 2021. Anion Exchange Resin Removal of Per- and Polyfluoroalkyl Substances (PFAS) from Impacted Water: A Critical Review. Water Research, 200, Article 117244. [https://doi.org/10.1016/j.watres.2021.117244 doi: 10.1016/j.watres.2021.117244]&nbsp;&nbsp; [[Media: BoyerEtAl2021a.pdf | Open Access Manuscript.pdf]]</ref>
  
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* ''In situ'' passive sampling of sediments in the Lower Duwamish Waterway Superfund site: Replicability, comparison with ''ex situ'' measurements, and use of data<ref name="Apell2016">Apell, J.N., and Gschwend, P.M., 2016. ''In situ'' passive sampling of sediments in the Lower Duwamish Waterway Superfund site: Replicability, comparison with ''ex situ'' measurements, and use of data. Environmental Pollution, 218, pp. 95-101. [https://doi.org/10.1016/j.envpol.2016.08.023 DOI: 10.1016/j.envpol.2016.08.023]&nbsp;&nbsp; [[Media: ApellGschwend2016.pdf | Authors’ Manuscript]]</ref>
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*Regenerable Resin Sorbent Technologies with Regenerant Solution Recycling for Sustainable Treatment of PFAS; SERDP Project ER18-1063 Final Report<ref>Strathmann, T.J., Higgins, C.P., Boyer, T., Schaefer, C., Ellis, A., Fang, Y., del Moral, L., Dietz, R., Kassar, C., Graham, C, 2023. Regenerable Resin Sorbent Technologies with Regenerant Solution Recycling for Sustainable Treatment of PFAS; SERDP Project ER18-1063 Final Report. 285 pages. [https://serdp-estcp.org/projects/details/d3ede38b-9f24-4b22-91c9-1ad634aa5384 Project Website]&nbsp;&nbsp; [[Media: ER18-1063.pdf | Report.pdf]]</ref>
  
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* Laboratory, Field, and Analytical Procedures for Using Passive Sampling in the Evaluation of Contaminated Sediments: User’s Manual<ref name="Burgess2017">Burgess, R.M., Kane Driscoll, S.B., Burton, A., Gschwend, P.M., Ghosh, U., Reible, D., Ahn, S., and Thompson, T., 2017. Laboratory, Field, and Analytical Procedures for Using Passive Sampling in the Evaluation of Contaminated Sediments: User’s Manual, EPA/600/R-16/357. SERDP/ESTCP and U.S. EPA, Office of Research and Development, Washington, DC 20460. [https://cfpub.epa.gov/si/si_public_record_report.cfm?Lab=NHEERL&dirEntryID=308731 Website]&nbsp;&nbsp; [[Media: EPA600R16357.pdf | Report.pdf]]</ref>
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==Introduction==
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[[File:StrathmannFig1.png | thumb |400px|Figure 1. Illustration of PFAS adsorption by anion exchange resins (AERs). Incorporation of longer alkyl group side chains on the cationic quaternary amine functional groups leads to PFAS-resin hydrophobic interactions that increase resin selectivity for PFAS over inorganic anions like Cl<sup>-</sup>.]]
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Anion exchange is an adsorptive treatment technology that uses polymeric resin beads (0.5–1 mm diameter) that incorporate cationic adsorption sites to remove anionic pollutants from water<ref>SenGupta, A.K., 2017. Ion Exchange in Environmental Processes: Fundamentals, Applications and Sustainable Technology. Wiley. ISBN:9781119157397  [https://onlinelibrary.wiley.com/doi/book/10.1002/9781119421252 Wiley Online Library]</ref>. Anions (e.g., NO<sub>3</sub><sup>-</sup>) are adsorbed by an ion exchange reaction with anions that are initially bound to the adsorption sites (e.g., Cl<sup>-</sup>) during resin preparation. Many per- and polyfluoroalkyl substances (PFAS) of concern, including [[Wikipedia: Perfluorooctanoic acid | perfluorooctanoic acid (PFOA)]] and [[Wikipedia: Perfluorooctanesulfonic acid | perfluorooctane sulfonate (PFOS)]], are present in contaminated water as anionic species that can be adsorbed by anion exchange reactions<ref name="BoyerEtAl2021a"/><ref name="DixitEtAl2021">Dixit, F., Dutta, R., Barbeau, B., Berube, P., Mohseni, M., 2021. PFAS Removal by Ion Exchange Resins: A Review. Chemosphere, 272, Article 129777. [https://doi.org/10.1016/j.chemosphere.2021.129777 doi: 10.1016/j.chemosphere.2021.129777]</ref>
 +
 
 +
==Advantages and Disadvantages==
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 +
===Advantages===
 +
In comparison to other reported PFAS destruction techniques, PRD offers several advantages:
 +
*Relative to UV/sodium sulfite and UV/sodium iodide systems, the fitted degradation rates in the micelle-accelerated PRD reaction system were ~18 and ~36 times higher, indicating the key role of the self-assembled micelle in creating a confined space for rapid PFAS destruction<ref name="ChenEtAl2020"/>. The negatively charged hydrated electron associated with the positively charged cetyltrimethylammonium ion (CTA<sup>+</sup>) forms the surfactant micelle to trap molecules with similar structures, selectively mineralizing compounds with both hydrophobic and hydrophilic groups (e.g., PFAS).
 +
*The PRD reaction does not require solid catalysts or electrodes, which can be expensive to acquire and difficult to regenerate or dispose.  
 +
*The aqueous solution is not heated or pressurized, and the UV wavelength used does not cause direct water [[Wikipedia: Photodissociation | photolysis]], therefore the energy input to the system is more directly employed to destroy PFAS, resulting in greater energy efficiency.
 +
*Since the reaction is performed at ambient temperature and pressure, there are limited concerns regarding environmental health and safety or volatilization of PFAS compared to heated and pressurized systems.  
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*Due to the reductive nature of the reaction, there is no formation of unwanted byproducts resulting from oxidative processes, such as [[Wikipedia: Perchlorate | perchlorate]]  generation during electrochemical oxidation<ref>Veciana, M., Bräunig, J., Farhat, A., Pype, M. L., Freguia, S., Carvalho, G., Keller, J., Ledezma, P., 2022. Electrochemical Oxidation Processes for PFAS Removal from Contaminated Water and Wastewater: Fundamentals, Gaps and Opportunities towards Practical Implementation. Journal of Hazardous Materials, 434, Article 128886. [https://doi.org/10.1016/j.jhazmat.2022.128886 doi: 10.1016/j.jhazmat.2022.128886]</ref><ref>Trojanowicz, M., Bojanowska-Czajka, A., Bartosiewicz, I., Kulisa, K., 2018. Advanced Oxidation/Reduction Processes Treatment for Aqueous Perfluorooctanoate (PFOA) and Perfluorooctanesulfonate (PFOS) – A Review of Recent Advances. Chemical Engineering Journal, 336, pp. 170–199. [https://doi.org/10.1016/j.cej.2017.10.153 doi: 10.1016/j.cej.2017.10.153]</ref><ref>Wanninayake, D.M., 2021. Comparison of Currently Available PFAS Remediation Technologies in Water: A Review. Journal of Environmental Management, 283, Article 111977. [https://doi.org/10.1016/j.jenvman.2021.111977 doi: 10.1016/j.jenvman.2021.111977]</ref>.
 +
*Aqueous fluoride ions are the primary end products of PRD, enabling real-time reaction monitoring with a fluoride [[Wikipedia: Ion-selective electrode | ion selective electrode (ISE)]], which is far less expensive and faster than relying on PFAS analytical data alone to monitor system performance.
 +
 
 +
===Disadvantages===
 +
*The CTAB additive is only partially consumed during the reaction, and although CTAB is not problematic when discharged to downstream treatment processes that incorporate aerobic digestors, CTAB can be toxic to surface waters and anaerobic digestors. Therefore, disposal options for treated solutions will need to be evaluated on a site-specific basis. Possible options include removal of CTAB from solution for reuse in subsequent PRD treatments, or implementation of an oxidation reaction to degrade CTAB.
 +
*The PRD reaction rate decreases in water matrices with high levels of total dissolved solids (TDS). It is hypothesized that in high TDS solutions (e.g., ion exchange still bottoms with TDS of 200,000 ppm), the presence of ionic species inhibits the association of the electron donor with the micelle, thus decreasing the reaction rate.
 +
*The PRD reaction rate decreases in water matrices with very low UV transmissivity. Low UV transmissivity (i.e., < 1 %) prevents the penetration of UV light into the solution, such that the utilization efficiency of UV light decreases.
  
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==Introduction==
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==State of the Art==
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Environmental media such as sediments typically contain many different materials or phases, including liquid solutions (e.g. water, [[Light Non-Aqueous Phase Liquids (LNAPLs)| nonaqueous phase liquids]]like spilled oils) and diverse solids (e.g., quartz, aluminosilicate clays, and combustion-derived soots).  Further, the chemical concentration in the porewater medium includes both molecules that are "truly dissolved" in the water and others that are associated with colloids in the porewater<ref name="Brownawell1986">Brownawell, B.J., and Farrington, J.W., 1986. Biogeochemistry of PCBs in interstitial waters of a coastal marine sediment. Geochimica et Cosmochimica Acta, 50(1), pp. 157-169.  [https://doi.org/10.1016/0016-7037(86)90061-X DOI: 10.1016/0016-7037(86)90061-X]&nbsp;&nbsp; Free download available from: [https://semspub.epa.gov/work/01/268631.pdf US EPA].</ref><ref name="Chin1992">Chin, Y.P., and Gschwend, P.M., 1992. Partitioning of Polycyclic Aromatic Hydrocarbons to Marine Porewater Organic Colloids. Environmental Science and Technology, 26(8), pp. 1621-1626.  [https://doi.org/10.1021/es00032a020 DOI: 10.1021/es00032a020]</ref><ref name="Achman1996">Achman, D.R., Brownawell, B.J., and Zhang, L., 1996. Exchange of Polychlorinated Biphenyls Between Sediment and Water in the Hudson River Estuary. Estuaries, 19(4), pp. 950-965.  [https://doi.org/10.2307/1352310 DOI: 10.2307/1352310]&nbsp;&nbsp; Free download available from: [https://www.academia.edu/download/55010335/135231020171114-2212-b93vic.pdf Academia.edu]</ref>. As a result, contaminant chemicals distribute among these diverse media (Figure 1) according to their affinity for each and the amount of each phase in the system<ref name="Gustafsson1996">Gustafsson, Ö., Haghseta, F., Chan, C., MacFarlane, J., and Gschwend, P.M., 1996. Quantification of the Dilute Sedimentary Soot Phase: Implications for PAH Speciation and Bioavailability. Environmental Science and Technology, 31(1), pp. 203-209.  [https://doi.org/10.1021/es960317s  DOI: 10.1021/es960317s]</ref><ref name="Luthy1997">Luthy, R.G., Aiken, G.R., Brusseau, M.L., Cunningham, S.D., Gschwend, P.M., Pignatello, J.J., Reinhard, M., Traina, S.J., Weber, W.J., and Westall, J.C., 1997. Sequestration of Hydrophobic Organic Contaminants by Geosorbents. Environmental Science and Technology, 31(12), pp. 3341-3347.  [https://doi.org/10.1021/es970512m DOI: 10.1021/es970512m]</ref><ref name="Lohmann2005">Lohmann, R., MacFarlane, J.K., and Gschwend, P.M., 2005. Importance of Black Carbon to Sorption of Native PAHs, PCBs, and PCDDs in Boston and New York Harbor Sediments. Environmental Science and Technology, 39(1), pp.141-148.  [https://doi.org/10.1021/es049424+  DOI: 10.1021/es049424+]</ref><ref name="Cornelissen2005">Cornelissen, G., Gustafsson, Ö., Bucheli, T.D., Jonker, M.T., Koelmans, A.A., and van Noort, P.C., 2005. Extensive Sorption of Organic Compounds to Black Carbon, Coal, and Kerogen in Sediments and Soils: Mechanisms and Consequences for Distribution, Bioaccumulation, and Biodegradation. Environmental Science and Technology, 39(18), pp. 6881-6895.  [https://doi.org/10.1021/es050191b  DOI: 10.1021/es050191b]</ref><ref name="Koelmans2009">Koelmans, A.A., Kaag, K., Sneekes, A., and Peeters, E.T.H.M., 2009. Triple Domain in Situ Sorption Modeling of Organochlorine Pesticides, Polychlorobiphenyls, Polyaromatic Hydrocarbons, Polychlorinated Dibenzo-p-Dioxins, and Polychlorinated Dibenzofurans in Aquatic Sediments. Environmental Science and Technology, 43(23), pp. 8847-8853.  [https://doi.org/10.1021/es9021188 DOI: 10.1021/es9021188]</ref>. As such, the chemical concentration in any one medium (e.g., truly dissolved in porewater) in a multi-material system like sediment is very hard to know from measures of the total sediment concentration, which unfortunately is the information typically found by analyzing for chemicals in sediment samples.
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 +
===Technical Performance===
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[[File:WittFig2.png | thumb |400px| Figure 2. Enspired Solutions<small><sup>TM</sup></small> commercial PRD PFAS destruction equipment, the PFASigator<small><sup>TM</sup></small>. Dimensions are 8 feet long by 4 feet wide by 9 feet tall.]]
  
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If an animal moves into this system, it will also accumulate the chemical in its tissues from the loads in all the other materials (Figure 1). This is important if one is concerned with exposures of the chemical to other organisms, including humans, who may eat such shellfish.  Predicting the quantity of contaminant in the clam requires knowledge of the relative affinities of the chemical for the clam versus the sediment materials.  For example, if one knew the chemical's truly dissolved concentration in the porewater and could reasonably assume the chemical of interest in the clams has mostly accumulated in its lipids (as is often the case for very hydrophobic compounds), then one could estimate the chemical concentration in the clam (''C<sub><small>clam</small></sub>'', typically in units of &mu;g/kg clam wet weight) using a lipid-water [[Wikipedia: Partition coefficient | partition coefficient]], ''K<sub><small>lipid-water</small></sub>'', typically in units of (&mu;g/kg lipid)'''/'''(&mu;g/L water), and the porewater concentration of the chemical (''C<sub><small>porewater</small></sub>'', in &mu;g/L) with Equation 1.
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{| class="wikitable mw-collapsible" style="float:left; margin-right:20px; text-align:center;"
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{|
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|+Table 1. Percent decreases from initial PFAS concentrations during benchtop testing of PRD treatment in different water matrices
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|
+
|-
 +
! Analytes
 +
!
 +
! GW
 +
! FF
 +
! AFFF<br>Rinsate
 +
! AFF<br>(diluted 10X)
 +
! IDW NF
 
|-
 
|-
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| || Equation 1.
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| &Sigma; Total PFAS<small><sup>a</sup></small> (ND=0)
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| style="text-align:center;"| <big>'''''C<sub><small>clam</small></sub> '''=''' f<sub><small>lipid</small></sub> '''x''' K<sub><small>lipid-water</small></sub> '''x''' C<sub><small>porewater</small></sub>'''''</big>
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| rowspan="9" style="background-color:white;" | <p style="writing-mode: vertical-rl">% Decrease<br>(Initial Concentration, &mu;g/L)</p>
 +
| 93%<br>(370) || 96%<br>(32,000) || 89%<br>(57,000) || 86 %<br>(770,000) || 84%<br>(82)
 
|-
 
|-
−
| where:
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| &Sigma; Total PFAS (ND=MDL) || 93%<br>(400) || 86%<br>(32,000) || 90%<br>(59,000) || 71%<br>(770,000) || 88%<br>(110)
 +
|-
 +
| &Sigma; Total PFAS (ND=RL) || 94%<br>(460) || 96%<br>(32,000) || 91%<br>(66,000) || 34%<br>(770,000) || 92%<br>(170)
 
|-
 
|-
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| || ''f<sub><small>lipid</small></sub>'' || is the fraction lipids contribute to the total wet weight of a clam (kg lipid/kg clam wet weight), and
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| &Sigma; Highly Regulated PFAS<small><sup>b</sup></small> (ND=0) || >99%<br>(180) || >99%<br>(20,000) || 95%<br>(20,000) || 92%<br>(390,000) || 95%<br>(50)
 
|-
 
|-
−
| || ''C<sub><small>porewater</small></sub>'' || is the freely dissolved contaminant concentration in the porewater surrounding the clam.
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| &Sigma; Highly Regulated PFAS (ND=MDL) || >99%<br>(180) || 98%<br>(20,000) || 95%<br>(20,000) || 88%<br>(390,000) || 95%<br> (52)
−
|}
 
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While there is a great deal of information on the values of ''K<sub><small>lipid-water</small></sub>'' for many chemicals<ref name="Schwarzenbach2017">Schwarzenbach, R.P., Gschwend, P.M., and Imboden, D.M., 2017.  Environmental Organic Chemistry, 3rd edition. Ch. 16: Equilibrium Partitioning from Water and Air to Biota, pp. 469-521. John Wiley and Sons.  ISBN: 978-1-118-76723-8</ref>, it is often very inaccurate to estimate truly dissolved porewater concentrations from total sediment concentrations using assumptions about the affinity of those chemicals for the solids in the system<ref name="Gustafsson1996"/>. Further, it is difficult to isolate porewater without colloids and/or measure the very low truly dissolved concentrations of hydrophobic contaminants of concern like [[Polycyclic Aromatic Hydrocarbons (PAHs) | polycyclic aromatic hydrocarbons (PAHs)]], [[Wikipedia: Polychlorinated biphenyl | polychlorinated biphenyls (PCBs)]], nonionic pesticides like [[Wikipedia: DDT | dichlorodiphenyltrichloroethane (DDT)]], and [[Wikipedia: Polychlorinated dibenzodioxins | polychlorinated dibenzo-p-dioxins (PCDDs)]]/[[Wikipedia: Polychlorinated dibenzofurans | dibenzofurans (PCDFs)]]<ref name="Hawthorne2005">Hawthorne, S.B., Grabanski, C.B., Miller, D.J., and Kreitinger, J.P., 2005. Solid-Phase Microextraction Measurement of Parent and Alkyl Polycyclic Aromatic Hydrocarbons in Milliliter Sediment Pore Water Samples and Determination of K<sub><small>DOC</small></sub> Values. Environmental Science and Technology, 39(8), pp. 2795-2803.  [https://doi.org/10.1021/es0405171 DOI: 10.1021/es0405171]</ref>.
 
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==Passive Samplers==
 
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One approach to address this problem for contaminated sediments is to insert organic polymers like low density polyethylene (LDPE), polydimethylsiloxane (PDMS), or polyoxymethylene (POM) that can absorb such chemicals in the sediment<ref name="Mayer2000">Mayer, P., Vaes, W.H., Wijnker, F., Legierse, K.C., Kraaij, R., Tolls, J., and Hermens, J.L., 2000. Sensing Dissolved Sediment Porewater Concentrations of Persistent and Bioaccumulative Pollutants Using Disposable Solid-Phase Microextraction Fibers. Environmental Science and Technology, 34(24), pp. 5177-5183.  [https://doi.org/10.1021/es001179g DOI: 10.1021/es001179g]</ref><ref name="Booij2003">Booij, K., Hoedemaker, J.R., and Bakker, J.F., 2003. Dissolved PCBs, PAHs, and HCB in Pore Waters and Overlying Waters of Contaminated Harbor Sediments. Environmental Science and Technology, 37(18), pp. 4213-4220.  [https://doi.org/10.1021/es034147c DOI: 10.1021/es034147c]</ref><ref name="Cornelissen2008">Cornelissen, G., Pettersen, A., Broman, D., Mayer, P., and Breedveld, G.D., 2008. Field testing of equilibrium passive samplers to determine freely dissolved native polycyclic aromatic hydrocarbon concentrations. Environmental Toxicology and Chemistry, 27(3), pp. 499-508.  [https://doi.org/10.1897/07-253.1 DOI: 10.1897/07-253.1]</ref><ref name="Tomaszewski2008">Tomaszewski, J.E., and Luthy, R.G., 2008. Field Deployment of Polyethylene Devices to Measure PCB Concentrations in Pore Water of Contaminated Sediment. Environmental Science and Technology, 42(16), pp. 6086-6091.  [https://doi.org/10.1021/es800582a DOI: 10.1021/es800582a]</ref><ref name="Fernandez2009">Fernandez, L.A., MacFarlane, J.K., Tcaciuc, A.P., and Gschwend, P.M., 2009. Measurement of Freely Dissolved PAH Concentrations in Sediment Beds Using Passive Sampling with Low-Density Polyethylene Strips. Environmental Science and Technology, 43(5), pp. 1430-1436.  [https://doi.org/10.1021/es802288w DOI: 10.1021/es802288w]</ref><ref name="Arp2015">Arp, H.P.H., Hale, S.E., Elmquist Kruså, M., Cornelissen, G., Grabanski, C.B., Miller, D.J., and Hawthorne, S.B., 2015. Review of polyoxymethylene passive sampling methods for quantifying freely dissolved porewater concentrations of hydrophobic organic contaminants. Environmental Toxicology and Chemistry, 34(4), pp. 710-720.  [https://doi.org/10.1002/etc.2864 DOI: 10.1002/etc.2864]&nbsp;&nbsp;  [https://setac.onlinelibrary.wiley.com/doi/epdf/10.1002/etc.2864 Free access article.]&nbsp;&nbsp; [[Media: Arp2015.pdf | Report.pdf]]</ref><ref name="Apell2016"/>. In this approach, the polymer is inserted in the sediment bed where it absorbs some of the contaminant load via the contaminant's diffusion into the polymer from the surroundings. When the polymer achieves sorptive equilibration with the sediments, the chemical concentration in the polymer, ''C<sub><small>polymer</small></sub>'' (&mu;g/kg polymer), can be used to find the corresponding concentration in the porewater,  ''C<sub><small>porewater</small></sub>'' (&mu;g/L), using a polymer-water partition coefficient, ''K<sub><small>polymer-water</small></sub>'' ((&mu;g/kg polymer)'''/'''(&mu;g/L water)), that has previously been found in laboratory testing<ref name="Lohmann2012">Lohmann, R., 2012. Critical Review of Low-Density Polyethylene’s Partitioning and Diffusion Coefficients for Trace Organic Contaminants and Implications for Its Use as a Passive Sampler. Environmental Science and Technology, 46(2), pp. 606-618.  [https://doi.org/10.1021/es202702y DOI: 10.1021/es202702y]</ref><ref name="Ghosh2014">Ghosh, U., Kane Driscoll, S., Burgess, R.M., Jonker, M.T., Reible, D., Gobas, F., Choi, Y., Apitz, S.E., Maruya, K.A., Gala, W.R., Mortimer, M., and Beegan, C., 2014. Passive Sampling Methods for Contaminated Sediments: Practical Guidance for Selection, Calibration, and Implementation. Integrated Environmental Assessment and Management, 10(2), pp. 210-223.  [https://doi.org/10.1002/ieam.1507 DOI: 10.1002/ieam.1507]&nbsp;&nbsp; [https://setac.onlinelibrary.wiley.com/doi/epdf/10.1002/ieam.1507 Free access article.]&nbsp;&nbsp; [[Media: Ghosh2014.pdf | Report.pdf]]</ref>, as shown in Equation 2.
 
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{|
 
−
|
 
 
|-
 
|-
−
|&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;|| Equation&nbsp;2.
+
| &Sigma; Highly Regulated PFAS (ND=RL) || >99%<br>(190) || 93%<br>(20,000) || 95%<br>(20,000) || 79%<br>(390,000) || 95%<br>(55)
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| style="width:600px; text-align:center;" | <big>'''''C<sub><small>porewater</small></sub> '''=''' C<sub><small>polymer</small></sub> '''/''' K<sub><small>polymer-water</small></sub>'''''</big>
 
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|}
 
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Such “passive uptake” by the polymer also reflects the availability of the chemicals for transport to adjacent systems (e.g., overlying surface waters) and for uptake into organisms (e.g., [[Wikipedia: Bioaccumulation | bioaccumulation]]).  Thus, one can use the porewater concentrations to estimate the biotic accumulation of the chemicals, too.  For example, for the concentration in the clam equilibrated with the sediment, ''C<sub><small>clam</small></sub>'' (&mu;g/kg clam), would be found by combining Equations 1 and 2 to get Equation 3.
 
−
{|
 
−
|
 
 
|-
 
|-
−
|&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;&nbsp;|| Equation&nbsp;3.
+
| &Sigma; High Priority PFAS<small><sup>c</sup></small> (ND=0) || 91%<br>(180) || 98%<br>(20,000) || 85%<br>(20,000) || 82%<br>(400,000) || 94%<br>(53)
−
|style="width:700px; text-align:center;" |<big>'''''C<sub><small>clam</small></sub> '''=''' f<sub><small>lipid</small></sub> '''x''' K<sub><small>lipid-water</small></sub> '''x''' C<sub><small>polymer</small></sub> '''/''' K<sub><small>polymer-water</small></sub>'''''</big>
 
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|}
 
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==Performance Reference Compounds (PRCs)==
 
−
Perhaps unsurprisingly, pollutants with low water solubility like PAHs, PCBs, etc. do not diffuse quickly through sediment beds.  As a result, their accumulation in polymeric materials in sediments can take a long time to achieve equilibration<ref name="Fernandez2009b">Fernandez, L. A., Harvey, C.F., and Gschwend, P.M., 2009. Using Performance Reference Compounds in Polyethylene Passive Samplers to Deduce Sediment Porewater Concentrations for Numerous Target Chemicals. Environmental Science and Technology, 43(23), pp. 8888-8894. [https://doi.org/10.1021/es901877a DOI: 10.1021/es901877a]</ref><ref name="Lampert2015">Lampert, D.J., Thomas, C., and Reible, D.D., 2015. Internal and external transport significance for predicting contaminant uptake rates in passive samplers. Chemosphere, 119, pp. 910-916.  [https://doi.org/10.1016/j.chemosphere.2014.08.063 DOI: 10.1016/j.chemosphere.2014.08.063]&nbsp;&nbsp; Free download available from: [https://www.academia.edu/download/44146586/chemosphere_2014.pdf Academia.edu]</ref><ref name="Apell2016b">Apell, J.N., Tcaciuc, A.P., and Gschwend, P.M., 2016. Understanding the rates of nonpolar organic chemical accumulation into passive samplers deployed in the environment: Guidance for passive sampler deployments. Integrated Environmental Assessment and Management, 12(3), pp. 486-492.  [https://doi.org/10.1002/ieam.1697 DOI: 10.1002/ieam.1697]</ref>. This problem was recognized previously for passive samplers called [[Wikipedia: Semipermeable membrane devices | semipermeable membrane devices]] (SPMDs, e.g. polyethylene bags filled with triolein<ref name="Huckins2002">Huckins, J.N., Petty, J.D., Lebo, J.A., Almeida, F.V., Booij, K., Alvarez, D.A., Cranor, W.L., Clark, R.C., and Mogensen, B.B., 2002. Development of the Permeability/Performance Reference Compound Approach for In Situ Calibration of Semipermeable Membrane Devices. Environmental Science and Technology, 36(1), pp. 85-91.  [https://doi.org/10.1021/es010991w DOI: 10.1021/es010991w]</ref>) that were deployed in surface waters. As a result, representative chemicals called performance reference compound (PRCs) were dosed inside the samplers before their deployment in the environment, and the PRCs' diffusive losses out of the SPMD could be used to quantify the fractional approach toward sampler-environmental surroundings equilibration<ref name="Booij2002">Booij, K., Smedes, F., and van Weerlee, E.M., 2002. Spiking of performance reference compounds in low density polyethylene and silicone passive water samplers. Chemosphere 46(8), pp.1157-1161.  [https://doi.org/10.1016/S0045-6535(01)00200-4 DOI: 10.1016/S0045-6535(01)00200-4]</ref><ref name="Huckins2002"/>. A similar approach can be used for polymers inserted in sediment beds<ref name="Fernandez2009b"/><ref name="Apell2014"/>. Commonly, isotopically labeled forms of the compounds of interest such as deuterated or <sup>13</sup>C-labelled PAHs or PCBs are homogeneously impregnated into the polymers before their deployments.  Upon insertion of the polymer into the sediment bed (or overlying waters or even air), the initially evenly distributed PRCs begin to diffuse out of the sampling polymer and  into the sediment (Figure 2).
 
−
 
 
−
Assuming the contaminants of interest undergo the same mass transfer restrictions limiting their rates of uptake into the polymer (e.g., diffusion through the sedimentary porous medium) that are also limiting transfers of the PRCs out of the polymer<ref name="Fernandez2009b"/><ref name="Apell2014"/>, then fractional losses of the PRCs during a particular deployment can be used to adjust the accumulated contaminant loads to what they would have been at equilibrium with their surroundings with Equation 4.
 
−
{|
 
−
|
 
 
|-
 
|-
−
| || Equation 4.
+
| &Sigma; High Priority PFAS (ND=MDL) || 91%<br>(190) || 94%<br>(20,000) || 85%<br>(20,000) || 79%<br>(400,000) || 86%<br>(58)
−
| style="text-align:center;"| <big>'''''C(&inf;)<sub><small>polymer</small></sub> '''=''' C(t)<sub><small>polymer</small></sub> '''/''' f<sub><small>PRC lost</small></sub>'''''</big>
 
 
|-
 
|-
−
| where:
+
| &Sigma; High Priority PFAS (ND=RL) || 92%<br>(200) || 87%<br>(20,000) || 86%<br>(21,000) || 70%<br>(400,000) || 87%<br>(65)
 
|-
 
|-
−
| || ''f<sub><small>PRC lost</small></sub>'' || is the fraction of the PRC lost to outward diffusion,
+
| Fluorine mass balance<small><sup>d</sup></small> || ||106% || 109% || 110% || 65% || 98%
 
|-
 
|-
−
| || ''C(&inf;)<sub><small>polymer</small></sub>'' || is the concentration of the contaminant in the polymer at equilibrium, and
+
| Sorbed organic fluorine<small><sup>e</sup></small> || || 4% || 4% || 33% || N/A || 31%
 
|-
 
|-
−
| || ''C(t)<sub><small>polymer</small></sub>'' || is the concentration of the contaminant in the polymer after deployment time, t.
+
| colspan="7" style="background-color:white; text-align:left" | <small>Notes:<br>GW = groundwater<br>GW FF = groundwater foam fractionate<br>AFFF rinsate = rinsate collected from fire system decontamination<br>AFFF (diluted 10x) = 3M Lightwater AFFF diluted 10x<br>IDW NF = investigation derived waste nanofiltrate<br>ND = non-detect<br>MDL = Method Detection Limit<br>RL = Reporting Limit<br><small><sup>a</sup></small>Total PFAS = 40 analytes + unidentified PFCA precursors<br><small><sup>b</sup></small>Highly regulated PFAS = PFNA, PFOA, PFOS, PFHxS, PFBS, HFPO-DA<br><small><sup>c</sup></small>High priority PFAS = PFNA, PFOA, PFHxA, PFBA, PFOS, PFHxS, PFBS, HFPO-DA<br><small><sup>d</sup></small>Ratio of the final to the initial organic fluorine plus inorganic fluoride concentrations<br><small><sup>e</sup></small>Percent of organic fluorine that sorbed to the reactor walls during treatment<br></small>
−
|}  
+
|}
−
 
+
</br>
−
 
+
The&nbsp;PRD&nbsp;reaction&nbsp;has&nbsp;been validated at the bench scale for the destruction of PFAS in a variety of environmental samples from Department of Defense sites (Table 1). Enspired Solutions<small><sup>TM</sup></small> has designed and manufactured a fully automatic commercial-scale piece of equipment called PFASigator<small><sup>TM</sup></small>, specializing in PRD PFAS destruction (Figure 2). This equipment is modular and scalable, has a small footprint, and can be used alone or in series with existing water treatment trains. The PFASigator<small><sup>TM</sup></small> employs commercially available UV reactors and monitoring meters that have been used in the water industry for decades. The system has been tested on PRD efficiency operational parameters, and key metrics were proven to be consistent with benchtop studies.
  
 +
Bench scale PRD tests were performed for the following samples collected from Department of Defense sites: groundwater (GW), groundwater foam fractionate (FF), firefighting truck rinsate ([[Wikipedia: Firefighting foam | AFFF]] Rinsate), 3M Lightwater AFFF, investigation derived waste nanofiltrate (IDW NF), [[Wikipedia: Ion exchange | ion exchange]] still bottom (IX SB), and Ansulite AFFF. The PRD treatment was more effective in low conductivity/TDS solutions. Generally, PRD reaction rates decrease for solutions with a TDS > 10,000 ppm, with an upper limit of 30,000 ppm. Ansulite AFFF and IX SB samples showed low destruction efficiencies during initial screening tests, which was primarily attributed to their high TDS concentrations. Benchtop testing data are shown in Table 1 for the remaining five sample matrices.
  
−
 
+
During treatment, PFOS and PFOA concentrations decreased 96% to >99% and 77% to 97%, respectively. For the PFAS with proposed drinking water Maximum Contaminant Levels (MCLs) recently established by the USEPA (PFNA, PFOA, PFOS, PFHxS, PFBS, and HFPO-DA), concentrations decreased >99% for GW, 93% for FF, 95% for AFFF Rinsate and IDW NF, and 79% for AFFF (diluted 10x) during the treatment time allotted. Meanwhile, the total PFAS concentrations, including all 40 known PFAS analytes and unidentified perfluorocarboxylic acid (PFCA) precursors, decreased from 34% to 96% following treatment. All of these concentration reduction values were calculated by using reporting limits (RL) as the concentrations for non-detects.  
−
[[File: Schwartz1w2Fig1.PNG | thumb | 500px | Figure 1.  Conceptual model of mercury speciation in the environment<ref>European Commission's Joint Research Centre, 2017. A new CRM to make mercury measurements in food more reliable. [https://ec.europa.eu/jrc/en/science-update/new-crm-make-mercury-measurements-food-more-reliable Website]</ref>]]
 
−
[[Wikipedia: Mercury (element) | Mercury]] (Hg) is released into the environment typically in the inorganic form. Natural emissions of Hg(0) come mainly from volcanoes and the ocean. Anthropogenic emissions are mainly from artisanal and small-scale gold mining, coal combustion, and various industrial processes that use Hg ( see the [https://www.unep.org/explore-topics/chemicals-waste/what-we-do/mercury/global-mercury-assessment UN Global mercury assessment]). Industrial and natural emissions of gaseous elemental mercury, Hg(0), can travel long distances in the atmosphere before being oxidized and deposited on land and in water as inorganic Hg(II). The long range transport and atmospheric deposition of Hg results in widespread low-level Hg contamination of soils at concentrations of 0.01 to 0.3 mg/kg<ref name="Eckley2020"/>.  
 
  
−
Hg-contaminated sites are most commonly contaminated with Hg(II) from industrial discharge and have soil concentrations in the range of 100s to 1000s of mg/kg<ref name="Eckley2020"/>. Direct exposure to Hg(II) and Hg(0) can be a human health risk at heavily contaminated sites. However, the organic form of Hg, [[Wikipedia: Methylmercury | methylmercury]] (MeHg or CH<sub>3</sub>Hg<sup>+</sup>) is typically the greater concern. MeHg is a neurotoxin that is particularly harmful to developing fetuses and young children. Direct contamination of the environment with MeHg is not common, but has occurred, most notably in [https://www.minamatadiseasemuseum.net/10-things-to-know Minamata Bay, Japan] (see also [https://en.wikipedia.org/wiki/Minamata_disease Minamata disease]). More commonly, MeHg is formed in the environment from Hg(II) in oxygen-limited conditions in a processes mediated by anaerobic microorganisms. Because MeHg [[Wikipedia: Biomagnification | biomagnifies]] in the aquatic food web, MeHg concentrations in fish can be elevated in areas that have relatively low levels of Hg contamination. The MeHg production depends heavily on site geochemistry, and high total Hg sediment concentrations do not always correlate with MeHg production potential.
+
Excellent fluorine/fluoride mass balance was achieved. There was nearly a 1:1 conversion of organic fluorine to free inorganic fluoride ion during treatment of GW, FF and AFFF Rinsate. The 3M Lightwater AFFF (diluted 10x) achieved only 65% fluorine mass balance, but this was likely due to high adsorption of PFAS to the reactor.
  
−
==Biogeochemistry/Mobility of Hg in soils==
+
===Application===
−
In the environment, Hg mobility is largely controlled by chelation with various ligands or adsorption to particles<ref name ="Hsu-Kim2018"/>. Hg(II) is most strongly attracted to the sulfur functional groups in dissolved organic matter (DOM) and to sulfur ligands. Over time, newly released Hg(II) “ages” and becomes less reactive to ligands and is less likely to be found in the dissolved phase. Legacy Hg(II) found in sediments and soils is more likely to be strongly adsorbed to the soil matrix and not very bioavailable compared to newly released Hg(II)<ref name ="Hsu-Kim2018"/>. MeHg has mobility tendencies similar to Hg, with DOM and sulfur ligands competing with each other to form complexes with MeHg<ref name="Loux2007">Loux, N.T., 2007. An assessment of thermodynamic reaction constants for simulating aqueous environmental monomethylmercury speciation. Chemical Speciation and Bioavailability, 19(4), pp.183-196.  [https://doi.org/10.3184/095422907X255947  DOI: 10.3184/095422907X255947]&nbsp;&nbsp; [https://www.tandfonline.com/doi/pdf/10.3184/095422907X255947?needAccess=true Free access article]&nbsp;&nbsp; [[Media: Loux2007.pdf | Report.pdf]]</ref>. However, unlike Hg-S complexes, MeHg-S does not have limited solubility.
+
Due to the first-order kinetics of PRD, destruction of PFAS is most energy efficient when paired with a pre-concentration technology, such as foam fractionation (FF), nanofiltration, reverse osmosis, or resin/carbon adsorption, that remove PFAS from water. Application of the PFASigator<small><sup>TM</sup></small> is therefore proposed as a part of a PFAS treatment train that includes a pre-concentration step.
  
−
The bioavailability of Hg(II) is one of the factors controlling MeHg production in the environment. MeHg production occurs in anoxic environments and is affected by: (1) the bioavailability of Hg(II) complexes to Hg-[[Wikipedia: Methylation | methylating]] microorganisms, (2) the activity of Hg-methylating microorganisms, and (3) the rate of biotic and abiotic [[Wikipedia: Demethylation | demethylation]]. MeHg is produced by anaerobic microorganisms that contain the ''hgcAB'' gene<ref name="Parks2013">Parks, J.M., Johs, A., Podar, M., Bridou, R. Hurt, R.A., Smith, S.D., Tomanicek, S.J., Qian, Y., Brown, S.D., Brandt, C.C., Palumbo, A.V., Smith, J.C., Wall, J.D., Elias, D.A., Liang, L., 2013. The Genetic Basis for Bacterial Mercury Methylation. Science, 339(6125), pp. 1332-1335.  [https://science.sciencemag.org/content/339/6125/1332 DOI: 10.1126/science.1230667]</ref>. These microorganisms are a diverse group and include, sulfate-reducing bacteria, iron-reducing bacteria, and methanogenic bacteria. Site geochemistry has a significant effect on MeHg production. Methylating microorganisms are sensitive to oxygen, and MeHg production occurs in oxygen-depleted or anaerobic zones in the environment, such as anoxic aquatic sediments, saturated soils, and biofilms with anoxic microenvironments<ref name="Bravo2020">Bravo, A.G., Cosio, C., 2020. Biotic formation of methylmercury: A bio–physico–chemical conundrum. Limnology and Oceanography, 65(5), pp. 1010-1027. [https://doi.org/10.1002/lno.11366 DOI: 10.1002/lno.11366]&nbsp;&nbsp; [https://aslopubs.onlinelibrary.wiley.com/doi/epdf/10.1002/lno.11366 Free Access Article]&nbsp;&nbsp; [[Media: Bravo2020.pdf | Report.pdf]]</ref>. The activity of methylating microorganisms can be impacted by redox conditions, the concentrations of organic carbon, and different electron acceptors (e.g. sulfate vs iron)<ref name="Bravo2020"/>. Overall, MeHg concentrations and production are impacted by demethylation as well. Demethylation can occur both abiotically and biotically and occurs at a much faster rate than methylation. The main routes of abiotic demethylation are photochemical reactions and demethylation catalyzed by reduced sulfur surfaces<ref name="Du2019">Du, H. Ma, M., Igarashi, Y., Wang, D., 2019. Biotic and Abiotic Degradation of Methylmercury in Aquatic Ecosystems: A Review. Bulletin of Environmental Contamination and Toxicology, 102 pp. 605-611. [https://doi.org/10.1007/s00128-018-2530-2 DOI: 10.1007/s00128-018-2530-2]</ref><ref name="Jonsson2016">Jonsson, S., Mazrui, N.M., Mason, R.P., 2016. Dimethylmercury Formation Mediated by Inorganic and Organic Reduced Sulfur Surfaces. Scientific Reports, 6, Article 27958.  [https://doi.org/10.1038/srep27958 DOI: 10.1038/srep27958]&nbsp;&nbsp; [https://www.nature.com/articles/srep27958.pdf Free access article]&nbsp;&nbsp; [[Media: Jonsson2016.pdf | Report.pdf]]</ref>. Methylmercury can be degraded biotically by aerobic bacteria containing the mercury detoxification, ''mer'' [[Wikipedia: Operon | operon]] and through oxidative demethylation by anaerobic microorganisms<ref name="Du2019"/>.
+
The first pilot study with the PFASigator<small><sup>TM</sup></small> was conducted in late 2023 at an industrial facility in Michigan with PFAS-impacted groundwater. The goal of the pilot study was to treat the groundwater to below the limits for regulatory discharge permits. For the pilot demonstration, the PFASigator<small><sup>TM</sup></small> was paired with an FF unit, which pre-concentrated the PFAS into a foamate that was pumped into the PFASigator<small><sup>TM</sup></small> for batch PFAS destruction. Residual PFAS remaining after the destruction batch was treated by looping back the PFASigator<small><sup>TM</sup></small> effluent to the FF system influent. During the one-month field pilot duration, site-specific discharge limits were met, and steady state operation between the FF unit and PFASigator<small><sup>TM</sup></small> was achieved such that the PFASigator<small><sup>TM</sup></small> destroyed the required concentrated PFAS mass and no off-site disposal of PFAS contaminated waste was required.
  
−
==Bioaccumulation and Toxicology==
 
−
Regulatory criteria are most often based on total Hg concentrations, however, MeHg is the form of Hg that can [[Wikipedia: Bioaccumulation | bioaccumulate]] in wildlife and is the greatest human and ecological health risk<ref name=”ATSDR1999”>Agency for Toxic Substances and Disease Registry (ATSDR), 1999. Toxicological Profile for Mercury.  [https://www.atsdr.cdc.gov/ToxProfiles/tp46.pdf Free download]&nbsp;&nbsp; [[Media: ATSDR1999.pdf | Report.pdf]]</ref>. MeHg represents over 95% of the Hg found in fish<ref name="Bloom1992">Bloom, N.S., 1992. On the Chemical Form of Mercury in Edible Fish and Marine Invertebrate Tissue. Canadian Journal of Fisheries and Aquatic Sciences 49(5), pp. 1010-117.  [https://doi.org/10.1139/f92-113 DOI: 10.1139/f92-113]</ref>. Hg and MeHg can be taken up directly from contaminated water into organisms, with the identity of the Hg-ligand complexes determining how readily the Hg is taken up into the organism<ref name="Kidd2012">Kidd, K., Clayden, M., Jardine, T., 2012. Bioaccumulation and Biomagnification of Mercury through Food Webs. Environmental Chemistry and Toxicology of Mercury, pp. 453-499. Liu, G., Yong, C. O’Driscoll, N., Eds. John Wiley and Sons, Inc. Hoboken, NJ.  [https://doi.org/10.1002/9781118146644.ch14 DOI: 10.1002/9781118146644.ch14]</ref>. Direct bioconcentration from water is the major uptake route at the base of the food web. Hg and MeHg can also enter the food web when benthic organisms ingest contaminated sediments<ref name="Mason2001">Mason, R.P., 2001. The Bioaccumulation of Mercury, Methylmercury and Other Toxic Elements into Pelagic and Benthic Organisms. Coastal and Estuarine Risk Assessment, pp. 127-149. Newman, M., Roberts, M., and Hale, R.C., Ed.s. CRC Press. ISBN: 978-1-4200-3245-1  Free download from: [https://www.researchgate.net/profile/Robert-Mason-13/publication/266354387_The_Bioaccumulation_of_Mercury_Methylmercury_and_Other_Toxic_Elements_into_Pelagic_and_Benthic_Organisms/links/55083eff0cf26ff55f80662d/The-Bioaccumulation-of-Mercury-Methylmercury-and-Other-Toxic-Elements-into-Pelagic-and-Benthic-Organisms.pdf ResearchGate]</ref>. Further up the food web organisms are exposed to Hg and MeHg both through exposure to contaminated water and through their diet. The higher up the trophic level, the more important dietary exposure becomes. Fish obtain more than 90% of Hg from their diet<ref name="Kidd2012"/>.
 
−
 
−
Humans are mainly exposed to Hg in the forms of MeHg and Hg(0). Hg(0) exposure comes from dental amalgams and industrial/contaminated site exposures. Hg(0) readily crosses the blood/brain barrier and mainly effects the nervous system and the kidneys<ref name="Clarkson2003">Clarkson, T.W., Magos, L., Myers, G.J., 2003. The Toxicology of Mercury — Current Exposures and Clinical Manifestations. New England Journal of Medicine, 349, pp. 1731-1737. [https://doi.org/10.1056/NEJMra022471 DOI: 10.1056/NEJMra022471]</ref>. MeHg exposure comes from the consumption of contaminated fish. In the human body, MeHg is readily absorbed through the gastrointestinal tract into the bloodstream and crosses the blood/brain barrier, affecting the central nervous system. MeHg can also pass through the placenta to the fetus and is particularly harmful to the developing nervous system of the fetus.
 
−
 
−
MeHg and Hg toxicity in the body occurs through multiple pathways and may be linked to the affinity of Hg for sulfur groups. Hg and MeHg bind to S-containing groups, which can block normal bodily functions<ref name="Bjørklund2017">Bjørklund, G., Dadar, M., Mutter, J. and Aaseth, J., 2017. The toxicology of mercury: Current research and emerging trends. Environmental Research, 159, pp.545-554.  [https://doi.org/10.1016/j.envres.2017.08.051 DOI: 10.1016/j.envres.2017.08.051]</ref>.
 
−
 
−
==Regulatory Framework for Mercury==
 
−
In the United States, mercury is regulated by several different [[Wikipedia: Mercury regulation in the United States | environmental laws]] including: the Mercury Export Ban Act of 2008, the Mercury-Containing and Rechargeable Battery Management Act of 1996, the Clean Air Act, the Clean Water Act, the Emergency Planning and Community Right-to-Know Act,  the Resource Conservation and Recovery Act, and the Safe Drinking Water Act<ref name=”USEPA2021”>US EPA, 2021.  Environmental Laws that Apply to Mercury.  [https://www.epa.gov/mercury/environmental-laws-apply-mercury US EPA Website]</ref>.
 
−
 
−
In 2013, the United States signed the international [https://www.epa.gov/international-cooperation/minamata-convention-mercury Minamata Convention on Mercury]. The Minamata Convention on Mercury seeks to address and reduce human activities that are contributing to widespread mercury pollution. Worldwide, 128 countries have signed the Convention.
 
−
 
−
==Remediation Technologies==
 
−
As a chemical element, Hg cannot be destroyed, so the goal of Hg-remediation is immobilization and prevention of food web bioaccumulation. At very highly contaminated sites (>100s ppm), sediments are often removed and landfilled<ref name="Eckley2020"/>. ''In situ'' capping is also a common remediation approach. Both dredging and capping can be costly and ecologically destructive, and the development of less invasive, less costly remediation technologies for Hg and MeHg contaminated sediments is an active research field. Eckley et al.<ref name="Eckley2020"/>and Wang et al.<ref name="Wang2020">Wang, L., Hou, D., Cao, Y., Ok, Y.S., Tack, F., Rinklebe, J., O’Connor, D., 2020. Remediation of mercury contaminated soil, water, and air: A review of emerging materials and innovative technologies. Environmental International, 134, 105281.  [https://doi.org/10.1016/j.envint.2019.105281  DOI: 10.1016/j.envint.2019.105281]&nbsp;&nbsp; [https://www.sciencedirect.com/science/article/pii/S0160412019324754 Free access article]</ref> give thorough reviews of standard and emerging technologies.
 
−
 
−
Recently application of ''in situ'' sorbents has garnered interest as a remediation solution for Hg<ref name="Eckley2020"/>. Many different materials, including biochar and various formulations of [[In Situ Treatment of Contaminated Sediments with Activated Carbon | activated carbon]], are successful in lowering porewater concentrations of Hg and MeHg in contaminated sediments<ref name="Gilmour2013">Gilmour, C.C., Riedel, G.S., Riedel, G., Kwon, S., Landis, R., Brown, S.S., Menzie, C.A., Ghosh, U., 2013. Activated Carbon Mitigates Mercury and Methylmercury Bioavailability in Contaminated Sediments. Environmental Science and Technology, 47(22), pp. 13001-13010.  [https://doi.org/10.1021/es4021074 DOI: 10.1021/es4021074]&nbsp;&nbsp; Free download from: [https://www.researchgate.net/profile/Steven-Brown-18/publication/258042399_Activated_Carbon_Mitigates_Mercury_and_Methylmercury_Bioavailability_in_Contaminated_Sediments/links/5702a10e08aea09bb1a30083/Activated-Carbon-Mitigates-Mercury-and-Methylmercury-Bioavailability-in-Contaminated-Sediments.pdf ResearchGate]</ref>. More research is needed to determine whether Hg and MeHg sorbed to these materials are available for uptake into organisms. Site biogeochemistry can also impact the efficacy of sorbent materials, with dissolved organic matter and sulfide concentrations impacting Hg and MeHg sorption. Overall, knowing site biogeochemical characteristics is important for predicting Hg mobility and MeHg production risks as well as for designing a remediation strategy that will be effective.
 
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<br clear="left" />
 
 
==References==
 
==References==
 
<references />
 
<references />
 +
 
==See Also==
 
==See Also==

Latest revision as of 20:09, 16 May 2024

PFAS Treatment by Anion Exchange

Anion exchange has emerged as one of the most effective and economical technologies for treatment of water contaminated by per- and polyfluoroalkyl substances (PFAS). Anion exchange resins (AERs) are polymer beads (0.5–1 mm diameter) incorporating cationic adsorption sites that attract anionic PFAS by a combination of electrostatic and hydrophobic mechanisms. Both regenerable and single-use resin treatment systems are being investigated, and results from pilot-scale studies show that AERs can treat much greater volumes of PFAS-contaminated water than comparable amounts of granular activated carbon (GAC) adsorbent media. Life cycle treatment costs and environmental impacts of anion exchange and other adsorbent technologies are highly dependent upon the treatment criteria selected by site managers to determine when media is exhausted and requires replacement or regeneration.

Related Article(s):

Contributor(s):

  • Dr. Timothy J. Strathmann
  • Dr. Anderson Ellis
  • Dr. Treavor H. Boyer

Key Resource(s):

  • Anion Exchange Resin Removal of Per- and Polyfluoroalkyl Substances (PFAS) from Impacted Water: A Critical Review[1]
  • Regenerable Resin Sorbent Technologies with Regenerant Solution Recycling for Sustainable Treatment of PFAS; SERDP Project ER18-1063 Final Report[2]

Introduction

File:StrathmannFig1.png
Figure 1. Illustration of PFAS adsorption by anion exchange resins (AERs). Incorporation of longer alkyl group side chains on the cationic quaternary amine functional groups leads to PFAS-resin hydrophobic interactions that increase resin selectivity for PFAS over inorganic anions like Cl-.

Anion exchange is an adsorptive treatment technology that uses polymeric resin beads (0.5–1 mm diameter) that incorporate cationic adsorption sites to remove anionic pollutants from water[3]. Anions (e.g., NO3-) are adsorbed by an ion exchange reaction with anions that are initially bound to the adsorption sites (e.g., Cl-) during resin preparation. Many per- and polyfluoroalkyl substances (PFAS) of concern, including perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS), are present in contaminated water as anionic species that can be adsorbed by anion exchange reactions[1][4]

Advantages and Disadvantages

Advantages

In comparison to other reported PFAS destruction techniques, PRD offers several advantages:

  • Relative to UV/sodium sulfite and UV/sodium iodide systems, the fitted degradation rates in the micelle-accelerated PRD reaction system were ~18 and ~36 times higher, indicating the key role of the self-assembled micelle in creating a confined space for rapid PFAS destruction[5]. The negatively charged hydrated electron associated with the positively charged cetyltrimethylammonium ion (CTA+) forms the surfactant micelle to trap molecules with similar structures, selectively mineralizing compounds with both hydrophobic and hydrophilic groups (e.g., PFAS).
  • The PRD reaction does not require solid catalysts or electrodes, which can be expensive to acquire and difficult to regenerate or dispose.
  • The aqueous solution is not heated or pressurized, and the UV wavelength used does not cause direct water photolysis, therefore the energy input to the system is more directly employed to destroy PFAS, resulting in greater energy efficiency.
  • Since the reaction is performed at ambient temperature and pressure, there are limited concerns regarding environmental health and safety or volatilization of PFAS compared to heated and pressurized systems.
  • Due to the reductive nature of the reaction, there is no formation of unwanted byproducts resulting from oxidative processes, such as perchlorate generation during electrochemical oxidation[6][7][8].
  • Aqueous fluoride ions are the primary end products of PRD, enabling real-time reaction monitoring with a fluoride ion selective electrode (ISE), which is far less expensive and faster than relying on PFAS analytical data alone to monitor system performance.

Disadvantages

  • The CTAB additive is only partially consumed during the reaction, and although CTAB is not problematic when discharged to downstream treatment processes that incorporate aerobic digestors, CTAB can be toxic to surface waters and anaerobic digestors. Therefore, disposal options for treated solutions will need to be evaluated on a site-specific basis. Possible options include removal of CTAB from solution for reuse in subsequent PRD treatments, or implementation of an oxidation reaction to degrade CTAB.
  • The PRD reaction rate decreases in water matrices with high levels of total dissolved solids (TDS). It is hypothesized that in high TDS solutions (e.g., ion exchange still bottoms with TDS of 200,000 ppm), the presence of ionic species inhibits the association of the electron donor with the micelle, thus decreasing the reaction rate.
  • The PRD reaction rate decreases in water matrices with very low UV transmissivity. Low UV transmissivity (i.e., < 1 %) prevents the penetration of UV light into the solution, such that the utilization efficiency of UV light decreases.

State of the Art

Technical Performance

Figure 2. Enspired SolutionsTM commercial PRD PFAS destruction equipment, the PFASigatorTM. Dimensions are 8 feet long by 4 feet wide by 9 feet tall.
Table 1. Percent decreases from initial PFAS concentrations during benchtop testing of PRD treatment in different water matrices
Analytes GW FF AFFF
Rinsate 		AFF
(diluted 10X) 		IDW NF
Σ Total PFASa (ND=0)

% Decrease
(Initial Concentration, μg/L)

 93%
(370)		 96%
(32,000)		 89%
(57,000)		 86 %
(770,000)		 84%
(82)
Σ Total PFAS (ND=MDL) 93%
(400)		 86%
(32,000)		 90%
(59,000)		 71%
(770,000)		 88%
(110)
Σ Total PFAS (ND=RL) 94%
(460)		 96%
(32,000)		 91%
(66,000)		 34%
(770,000)		 92%
(170)
Σ Highly Regulated PFASb (ND=0) >99%
(180)		 >99%
(20,000)		 95%
(20,000)		 92%
(390,000)		 95%
(50)
Σ Highly Regulated PFAS (ND=MDL) >99%
(180)		 98%
(20,000)		 95%
(20,000)		 88%
(390,000)		 95%
(52)
Σ Highly Regulated PFAS (ND=RL) >99%
(190)		 93%
(20,000)		 95%
(20,000)		 79%
(390,000)		 95%
(55)
Σ High Priority PFASc (ND=0) 91%
(180)		 98%
(20,000)		 85%
(20,000)		 82%
(400,000)		 94%
(53)
Σ High Priority PFAS (ND=MDL) 91%
(190)		 94%
(20,000)		 85%
(20,000)		 79%
(400,000)		 86%
(58)
Σ High Priority PFAS (ND=RL) 92%
(200)		 87%
(20,000)		 86%
(21,000)		 70%
(400,000)		 87%
(65)
Fluorine mass balanced 106% 109% 110% 65% 98%
Sorbed organic fluorinee 4% 4% 33% N/A 31%
Notes:
GW = groundwater
GW FF = groundwater foam fractionate
AFFF rinsate = rinsate collected from fire system decontamination
AFFF (diluted 10x) = 3M Lightwater AFFF diluted 10x
IDW NF = investigation derived waste nanofiltrate
ND = non-detect
MDL = Method Detection Limit
RL = Reporting Limit
aTotal PFAS = 40 analytes + unidentified PFCA precursors
bHighly regulated PFAS = PFNA, PFOA, PFOS, PFHxS, PFBS, HFPO-DA
cHigh priority PFAS = PFNA, PFOA, PFHxA, PFBA, PFOS, PFHxS, PFBS, HFPO-DA
dRatio of the final to the initial organic fluorine plus inorganic fluoride concentrations
ePercent of organic fluorine that sorbed to the reactor walls during treatment


The PRD reaction has been validated at the bench scale for the destruction of PFAS in a variety of environmental samples from Department of Defense sites (Table 1). Enspired SolutionsTM has designed and manufactured a fully automatic commercial-scale piece of equipment called PFASigatorTM, specializing in PRD PFAS destruction (Figure 2). This equipment is modular and scalable, has a small footprint, and can be used alone or in series with existing water treatment trains. The PFASigatorTM employs commercially available UV reactors and monitoring meters that have been used in the water industry for decades. The system has been tested on PRD efficiency operational parameters, and key metrics were proven to be consistent with benchtop studies.

Bench scale PRD tests were performed for the following samples collected from Department of Defense sites: groundwater (GW), groundwater foam fractionate (FF), firefighting truck rinsate ( AFFF Rinsate), 3M Lightwater AFFF, investigation derived waste nanofiltrate (IDW NF), ion exchange still bottom (IX SB), and Ansulite AFFF. The PRD treatment was more effective in low conductivity/TDS solutions. Generally, PRD reaction rates decrease for solutions with a TDS > 10,000 ppm, with an upper limit of 30,000 ppm. Ansulite AFFF and IX SB samples showed low destruction efficiencies during initial screening tests, which was primarily attributed to their high TDS concentrations. Benchtop testing data are shown in Table 1 for the remaining five sample matrices.

During treatment, PFOS and PFOA concentrations decreased 96% to >99% and 77% to 97%, respectively. For the PFAS with proposed drinking water Maximum Contaminant Levels (MCLs) recently established by the USEPA (PFNA, PFOA, PFOS, PFHxS, PFBS, and HFPO-DA), concentrations decreased >99% for GW, 93% for FF, 95% for AFFF Rinsate and IDW NF, and 79% for AFFF (diluted 10x) during the treatment time allotted. Meanwhile, the total PFAS concentrations, including all 40 known PFAS analytes and unidentified perfluorocarboxylic acid (PFCA) precursors, decreased from 34% to 96% following treatment. All of these concentration reduction values were calculated by using reporting limits (RL) as the concentrations for non-detects.

Excellent fluorine/fluoride mass balance was achieved. There was nearly a 1:1 conversion of organic fluorine to free inorganic fluoride ion during treatment of GW, FF and AFFF Rinsate. The 3M Lightwater AFFF (diluted 10x) achieved only 65% fluorine mass balance, but this was likely due to high adsorption of PFAS to the reactor.

Application

Due to the first-order kinetics of PRD, destruction of PFAS is most energy efficient when paired with a pre-concentration technology, such as foam fractionation (FF), nanofiltration, reverse osmosis, or resin/carbon adsorption, that remove PFAS from water. Application of the PFASigatorTM is therefore proposed as a part of a PFAS treatment train that includes a pre-concentration step.

The first pilot study with the PFASigatorTM was conducted in late 2023 at an industrial facility in Michigan with PFAS-impacted groundwater. The goal of the pilot study was to treat the groundwater to below the limits for regulatory discharge permits. For the pilot demonstration, the PFASigatorTM was paired with an FF unit, which pre-concentrated the PFAS into a foamate that was pumped into the PFASigatorTM for batch PFAS destruction. Residual PFAS remaining after the destruction batch was treated by looping back the PFASigatorTM effluent to the FF system influent. During the one-month field pilot duration, site-specific discharge limits were met, and steady state operation between the FF unit and PFASigatorTM was achieved such that the PFASigatorTM destroyed the required concentrated PFAS mass and no off-site disposal of PFAS contaminated waste was required.

References

  1. ^ 1.0 1.1 Boyer, T.H., Fang, Y., Ellis, A., Dietz, R., Choi, Y.J., Schaefer, C.E., Higgins, C.P., Strathmann, T.J., 2021. Anion Exchange Resin Removal of Per- and Polyfluoroalkyl Substances (PFAS) from Impacted Water: A Critical Review. Water Research, 200, Article 117244. doi: 10.1016/j.watres.2021.117244   Open Access Manuscript.pdf
  2. ^ Strathmann, T.J., Higgins, C.P., Boyer, T., Schaefer, C., Ellis, A., Fang, Y., del Moral, L., Dietz, R., Kassar, C., Graham, C, 2023. Regenerable Resin Sorbent Technologies with Regenerant Solution Recycling for Sustainable Treatment of PFAS; SERDP Project ER18-1063 Final Report. 285 pages. Project Website   Report.pdf
  3. ^ SenGupta, A.K., 2017. Ion Exchange in Environmental Processes: Fundamentals, Applications and Sustainable Technology. Wiley. ISBN:9781119157397 Wiley Online Library
  4. ^ Dixit, F., Dutta, R., Barbeau, B., Berube, P., Mohseni, M., 2021. PFAS Removal by Ion Exchange Resins: A Review. Chemosphere, 272, Article 129777. doi: 10.1016/j.chemosphere.2021.129777
  5. ^ Cite error: Invalid <ref> tag; no text was provided for refs named ChenEtAl2020
  6. ^ Veciana, M., Bräunig, J., Farhat, A., Pype, M. L., Freguia, S., Carvalho, G., Keller, J., Ledezma, P., 2022. Electrochemical Oxidation Processes for PFAS Removal from Contaminated Water and Wastewater: Fundamentals, Gaps and Opportunities towards Practical Implementation. Journal of Hazardous Materials, 434, Article 128886. doi: 10.1016/j.jhazmat.2022.128886
  7. ^ Trojanowicz, M., Bojanowska-Czajka, A., Bartosiewicz, I., Kulisa, K., 2018. Advanced Oxidation/Reduction Processes Treatment for Aqueous Perfluorooctanoate (PFOA) and Perfluorooctanesulfonate (PFOS) – A Review of Recent Advances. Chemical Engineering Journal, 336, pp. 170–199. doi: 10.1016/j.cej.2017.10.153
  8. ^ Wanninayake, D.M., 2021. Comparison of Currently Available PFAS Remediation Technologies in Water: A Review. Journal of Environmental Management, 283, Article 111977. doi: 10.1016/j.jenvman.2021.111977

See Also

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