A38. 20240819 brief OP aan gemeenteraden PFAS schade.pdf

There are, however, many wider benefits (or costs) that may result from remediation such as the loss or gain of soil functionality, which is difficult to account for in a CBA but is an important aspect of soil remediation (Chen and Li, 2018). The improvement of soil functionality and increased provision of ecosystem services could be an important benefit of using GRO that is currently neglected. Multi-criteria decision analysis is increasingly being used for evaluating positive and negative effects of remedial actions in the three domains of sustainability (environmental, social and economic) to support the decision on the most reasonable alternative taking into consideration other values which are not accounted for in a CBA (Bardos, 2014; Rosén et al., 2015; Söderqvist et al., 2015). Assumptions were made regarding the technical applications of certain remediation technologies that impacted the CBA results. For soil washing, it was assumed that 100 % of the excavated soil was washed and reused on site, which may not be completely accurate but is favourable based on the site geology of primarily sand. Similarly, thermal treatment is assumed for the treatment of fine aggregates with the reuse of coarse aggregates on-site, but the particle size distribution in terms of proportion of fine aggregates is unknown.

Also, the annual cost of phytoremediation was not possible to accurately quantify based on literature and could have affected the resulting NPV of Alt 3–5. The lump sum used as a present value of this cost includes total project costs (e.g., costs of establishment, operation and maintenance, biomass harvest and management) was consequently not sensitive to changes in the time required for risk reduction which could impact the present value of costs and benefits when discounted over a longer time. It is also a simplification to assume that full risk reduction will be achieved so quickly and in similar time for each hotspot remediation alternative. Other potentially viable PFAS remediation techniques such as engineered caps/covers were not included in this study but may be considered as similar to the considered containment techniques (i.e., S/S and stabilization with AC) and entail similar processes, costs, and limitations.

Funding This work is supported by Research Council Formas (Dnr 2018-01467; Dnr 2021-01428), a Swedish government research council for sustainable development. CRediT authorship contribution statement The authors contributed as follows: Conceptualization: P.D., Y.V., T.S., L.R., J.N.; Methodology: P.D., Y.V., T.S., L.R., J.N.; Validation: L.A.; Formal analysis: P.D., Y.V.; Investigation: P.D., Y.V.; Visualization: P.D., Y.V.; Writing - original draft preparation: P.D., Y.V., L.A., J.N.; Writing - review and editing: P.D., Y.V., L.A., T.S., L.R., J.N.; Supervision: Y.V., L.R., J.N.; Project administration: J.N.; Funding acquisition, Y.V., L.R., J.N. All authors have read and agreed to the published version of the manuscript. Data availability Data will be made available on request. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2023.163664.

6. Conclusions

References

The following main conclusions can be drawn from this study: • There are many uncertainties associated with PFAS contamination, including costs, benefits and effectiveness of PFAS remediation alternatives. Limitations in data and the novelty of PFAS remediation required making assumptions to compensate. Probabilistic CBA is demonstrated to be a robust method to account for uncertainties and parameter sensitivity and the model was further improved by creating multiple scenarios to test different model assumptions. The analysis provides valuable decision-support by evaluating the social profitability of different PFAS remediation alternatives.

Ahrens, L., Norström, K., Viktor, T., Cousins, A.P., Josefsson, S., 2015. Stockholm arlanda airport as a source of per- and polyfluoroalkyl substances to water, sediment and fish. Chemosphere 129, 33–38. https://doi.org/10.1016/j.chemosphere.2014.03.136. ATSDR, 2021. Toxicological Profile for Perfluoroalkyls. Bardos, P., 2014. Progress in sustainable remediation. Remediation 25, 23–32. https://doi. org/10.1002/rem.21412. Bardos, P., Spencer, K.L., Ward, R.D., Maco, B.H., Cundy, A.B., 2020. Integrated and sustainable management of post-industrial coasts. Front. Environ. Sci. 8, 1–14. https://doi.org/ 10.3389/fenvs.2020.00086. Bedford, T., Cooke, R., 2001. Probabilistic Risk Analysis: Foundations and Methods. Cambridge University Press, New York. 10

P. Drenning et al.

Science of the Total Environment 882 (2023) 163664 Mahinroosta, R., Senevirathna, L., 2020. A review of the emerging treatment technologies for PFAS contaminated soils. J. Environ. Manag. 255, 109896. https://doi.org/10.1016/j. jenvman.2019.109896. Merino, N., Qu, Y., Deeb, R.A., Hawley, E.L., Hoffmann, M.R., Mahendra, S., 2016. Degradation and removal methods for perfluoroalkyl and polyfluoroalkyl substances in water. Environ. Eng. Sci. 33, 615–649. https://doi.org/10.1089/ees.2016.0233. Ojo, A.F., Peng, C., Ng, J.C., 2021. Assessing the human health risks of per- and polyfluoroalkyl substances: a need for greater focus on their interactions as mixtures. J. Hazard. Mater. 407. https://doi.org/10.1016/j.jhazmat.2020.124863. Ok, Y.S., Rinklebe, J., Hou, D., Tsang, D.C.W., Tack, F.M.G.G., 2020. Soil and Groundwater Remediation Technologies: A Practical Guide. First edit. CRC Press, Boca Raton. Onwubuya, K., Cundy, A., Puschenreiter, M., Kumpiene, J., Bone, B., Greaves, J., Teasdale, P., Mench, M., Tlustos, P., Mikhalovsky, S., Waite, S., Friesl-Hanl, W., Marschner, B., Müller, I., 2009. Developing decision support tools for the selection of “gentle” remediation approaches. Sci. Total Environ. 407, 6132–6142. https://doi.org/10.1016/j.scitotenv. 2009.08.017. Pettersson, M., Ländell, M., Ohlsson, Y., Kleja, D.B., Tiberg, C., 2015.

Preliminära riktvärden för högfluorerade ämnen (PFAS) i mark och grundvatten, SGI. Rosén, L., Back, P.E., Söderqvist, T., Norrman, J., Brinkhoff, P., Norberg, T., Volchko, Y., Norin, M., Bergknut, M., Döberl, G., 2015. SCORE: a novel multi-criteria decision analysis approach to assessing the sustainability of contaminated land remediation. Sci. Total Environ. 511, 621–638. https://doi.org/10.1016/j.scitotenv.2014.12.058. Rosenqvist, L., Vestergren, R., Westberg, E., Eliaeson, K., Norström, K., Liljeberg, M., Strandberg, J., Rahmberg, M., 2017. Spridning av högfluorerade ämnen i mark från Stockholm Arlanda Airport: Förutsättningar för beräkning av platsspecifika riktvärden för mark. Ross, I., McDonough, J., Miles, J., Storch, P., Thelakkat Kochunarayanan, P., Kalve, E., Hurst, J., Dasgupta, S.S., Burdick, J., 2018. A review of emerging technologies for remediation of PFASs. Remediation 28, 101–126. https://doi.org/10.1002/rem.21553. SEPA, 2019. Vägledning om att riskbedöma och åtgärda PFAS-föroreningar inom förorenade områden, Rapport 6871 (In English: Guidance for Risk Assessing and Remediating PFAS Contamination at Contaminated Sites). Stockholm. Smith, J.W.N., Beuthe, B., Dunk, M., Demeure, S., Carmona, J.M.M., Medve, A., Spence, M.J., Pancras, T., Schrauwen, G., Held, T., Baker, K., Ross, I., Slenders, H., 2016.

Environmental fate and effects of polyand perfluoroalkyl substances (PFAS). CONCAWE Reports 1–107. Söderqvist, T., Brinkhoff, P., Norberg, T., Rosén, L., Back, P.E., Norrman, J., 2015. Cost-benefit analysis as a part of sustainability assessment ofremediation alternatives for contaminated land. J. Environ. Manag. 157, 267–278. https://doi.org/10.1016/j.jenvman. 2015.04.024. Sörengård, M., Gago-Ferrero, P., Kleja, D.B., Ahrens, L., 2021. Laboratory-scale and pilot-scale stabilization and solidification (S/S) remediation of soil contaminated with per- and polyfluoroalkyl substances (PFASs). J. Hazard. Mater., 402 https://doi.org/10.1016/j. jhazmat.2020.123453. STA, 2020. Analysmetod och samhällsekonomiska kalkylvärden för transportsektorn: ASEK 7.0. Swedish Traffic Administration. Volchko, Y., Norrman, J., Rosén, L., Karlfeldt Fedje, K., 2017. Cost-benefit analysis of copper recovery in remediation projects: a case study from Sweden. Sci. Total Environ. 605–606, 300–314. https://doi.org/10.1016/j.scitotenv.2017.06.128. Volchko, Y., Berggren Kleja, D., Back, P.E., Tiberg, C., Enell, A., Larsson, M., Jones, C.M., Taylor, A., Viketoft, M., Åberg, A., Dahlberg, A.K., Weiss, J., Wiberg, K., Rosén, L., 2020. Assessing costs and benefits of improved soil quality management in remediation projects: a study of an urban site contaminated with PAH and metals. Sci. Total Environ. 707.

https://doi.org/10.1016/j.scitotenv.2019.135582. Wan, X., Lei, M., Chen, T., 2016. Cost–benefit calculation of phytoremediation technology for heavy-metal-contaminated soil. Sci. Total Environ. 563–564, 796–802. https://doi.org/ 10.1016/j.scitotenv.2015.12.080. Wang, F., Lu, X., Li, X.Y., Shih, K., 2015. Effectiveness and mechanisms of defluorination of perfluorinated alkyl substances by calcium compounds during waste thermal treatment. Environ. Sci. Technol. 49, 5672–5680. https://doi.org/10.1021/es506234b. Wang, Yu., Chang, W., Wang, L., Zhang, Yinfeng, Zhang, Yuan, Wang, M., Wang, Yin, Li, P., 2019. A review of sources, multimedia distribution and health risks of novel fluorinated alternatives. Ecotoxicol. Environ. Saf. https://doi.org/10.1016/j.ecoenv.2019.109402. Wang, J., Lin, Z., He, X., Song, M., Westerhoff, P., Doudrick, K., Hanigan, D., 2022. Critical review of thermal decomposition of per- and polyfluoroalkyl substances: mechanisms and implications for thermal treatment processes. Environ. Sci. Technol. https://doi.org/10. 1021/acs.est.2c02251. Xu, Y., Hansson, E., Andersson, E.M., Jakobsson, K., Li, H., 2022. High exposure to perfluoroalkyl substances in drinking water is associated with increased risk of osteoporotic fractures – a cohort study from ronneby, Sweden. Environ. Res., 114796 https:// doi.org/10.1016/j.envres.2022.114796.

Bell, R., 1996. Quantifying diminution in value due to detrimental conditions: an application to environmentally contaminated properties. Environ. Claims J. 9, 127–137. Birgersson, L., Jouve, J., Jönsson, E., Asker, N., Andreasson, F., Golovko, O., Ahrens, L., Sturve, J., 2021. Thyroid function and immune status in perch (Perca fluviatilis) from lakes contaminated with PFASs or PCBs. Ecotoxicol. Environ. Saf. 222. https://doi.org/ 10.1016/j.ecoenv.2021.112495. Bolan, N., Sarkar, B., Yan, Y., Li, Q., Wijesekara, H., Kannan, K., Tsang, D.C.W., Schauerte, M., Bosch, J., Noll, H., Ok, Y.S., Scheckel, K., Kumpiene, J., Gobindlal, K., Kah, M., Sperry, J., Kirkham, M.B., Wang, H., Tsang, Y.F., Hou, D., Rinklebe, J., 2021. Remediation of polyand perfluoroalkyl substances (PFAS) contaminated soils – to mobilize or to immobilize or to degrade? J. Hazard. Mater. 401. https://doi.org/10.1016/j.jhazmat.2020.123892. Cao, L., Xu, W., Wan, Z., Li, G., Zhang, F., 2022. Occurrence of PFASs and its effect on soil bacteria at a fire-training area using PFOS-restricted aqueous film-forming foams. iScience, 25 https://doi.org/10.1016/j.isci.2022.104084. Chen, W., Li, H., 2018. Cost-effectiveness analysis for soil heavy metal contamination treatments. Water Air Soil Pollut. 229. https://doi.org/10.1007/s11270-018-3784-3. Cousins, I.T., Vestergren, R., Wang, Z., Scheringer, M., McLachlan, M.S., 2016.

The precautionary principle and chemicals management: the example of perfluoroalkyl acids in groundwater. Environ. Int. https://doi.org/10.1016/j.envint.2016.04.044. Cundy, A.B., Bardos, R.P., Puschenreiter, M., Mench, M., Bert, V., Friesl-Hanl, W., Müller, I., Li, X.N., Weyens, N., Witters, N., Vangronsveld, J., 2016. Brownfields to green fields: realising wider benefits from practical contaminant phytomanagement strategies. J. Environ. Manag. 184, 67–77. https://doi.org/10.1016/j.jenvman.2016.03.028. Drenning, P., Chowdhury, S., Volchko, Y., Rosén, L., Andersson-Sköld, Y., Norrman, J., 2022. A risk management framework for Gentle Remediation Options (GRO). Sci. Total Environ. 802, 149880. https://doi.org/10.1016/j.scitotenv.2021.149880. Evangelou, M.W.H., Robinson, B.H., 2022. The phytomanagement of PFAS-contaminated land. Int. J. Environ. Res. Public Health https://doi.org/10.3390/ijerph19116817. Fenton, S.E., Ducatman, A., Boobis, A., DeWitt, J.C., Lau, C., Ng, C., Smith, J.S., Roberts, S.M., 2021. Per- and Polyfluoroalkyl substance toxicity and human health review: current state of knowledge and strategies for informing future research. Environ. Toxicol. Chem. https://doi.org/10.1002/etc.4890. Franke, V., Ullberg, M., McCleaf, P., Wålinder, M., Köhler, S.J., Ahrens, L., 2021.

The Price of really clean water: combining nanofiltration with granular activated carbon and anion exchange resins for the removal of per- and polyfluoralkyl substances (PFASs) in drinking water production. ACS ES&T Water 1, 782–795. https://doi.org/10.1021/acsestwater. 0c00141. Gobelius, L., Lewis, J., Ahrens, L., 2017. Plant uptake of per- and Polyfluoroalkyl substances at a contaminated fire training facility to evaluate the phytoremediation potential of various plant species. Environ. Sci. Technol. 51, 12602–12610. https://doi.org/10.1021/acs.est. 7b02926. Goldenman, G., Fernandes, M., Holland, M., Tugran, T., Nordin, A., Schoumacher, C., McNeill, A., 2019. The Cost of Inaction: A Socioeconomic Analysis of Environmental and Health Impacts Linked to Exposure to PFAS, TemaNord. https://doi.org/10.21557/dsp. 51953910. Grimison, C., Knight, E.R., Nguyen, T.M.H., Nagle, N., Kabiri, S., Bräunig, J., Navarro, D.A., Kookana, R.S., Higgins, C.P., McLaughlin, M.J., Mueller, J.F., 2023. The efficacy of soil washing for the remediation of per- and poly-fluoroalkyl substances (PFASs) in the field. J. Hazard. Mater. 445. https://doi.org/10.1016/j.jhazmat.2022.130441. Harwell, M.C., Jackson, C., Kravitz, M., Lynch, K., Tomasula, J., Neale, A., Mahoney, M., Pachon, C., Scheuermann, K., Grissom, G., Parry, K., 2021.

Ecosystem services consideration in the remediation process for contaminated sites. J. Environ. Manag. 285, 112102. https://doi.org/10.1016/j.jenvman.2021.112102. Held, T., Reinhard, M., 2020. Remediation management for local and wide-spread PFAS contaminations (TEXTE 205/2020). Umwelt Bundesamt 310. Huff, D.K., Morris, L.A., Sutter, L., Costanza, J., Pennell, K.D., 2020. Accumulation of six PFAS compounds by woody and herbaceous plants: potential for phytoextraction. Int J Phytoremediation 22, 1538–1550. https://doi.org/10.1080/15226514.2020.1786004. ITRC, 2018. Remediation Technologies and Methods for Per- and Polyfluoroalkyl Substances (PFAS). ITRC PFAS Fact Sheets, pp. 1–24. ITRC, 2020a. History and Use of Per- and Polyfluoroalkyl Substances (PFAS) Found in the Environment. ITRC PFAS Fact Sheets 2000, pp. 1–4. ITRC, 2020b. Fate and Transport of Per- and Polyfluoroalkyl Substances (PFAS). ITRC PFAS Fact Sheets 1, pp. 1–4. Johansson, P.-O., Kriström, B., 2018. Cost-Benefit Analysis, Cambridge. Cambridge University Press, Cambridge. Lee, J.W., Choi, K., Park, K., Seong, C., Yu, S.do, Kim, P., 2020. Adverse effects of perfluoroalkyl acids on fish and other aquatic organisms: a review. Sci. Total Environ. https://doi.org/10.1016/j.scitotenv.2019.135334. Linacre, N.A., Whiting, S.N., Angle, J.S., 2005. The impact of uncertainty on phytoremediation project costs. Int. J.

Phytoremediation 7, 259–269. https://doi.org/ 10.1080/16226510500327103. Lu, D., Sha, S., Luo, J., Huang, Z., Zhang Jackie, X., 2020. Treatment train approaches for the remediation of per- and polyfluoroalkyl substances (PFAS): a critical review. J. Hazard. Mater. 386, 121963. https://doi.org/10.1016/j.jhazmat.2019.121963.

11

Update Science of the Total Environment Volume 916, Issue , 15 March 2024, Page DOI: https://doi.org/10.1016/j.scitotenv.2024.170429

Science of the Total Environment 916 (2024) 170429

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Corrigendum

Corrigendum to “Comparison of PFAS soil remediation alternatives at a civilian airport using cost-benefit analysis” [Sci. Total Environ. 882 (2023) 163664] P. Drenning a, *, Y. Volchko a, L. Ahrens b, L. Rosén a, T. Söderqvist c, J. Norrman a a b c

Department of Architecture and Civil Engineering, Chalmers University of Technology, SE-41296 Gothenburg, Sweden Department of Aquatic Sciences and Assessment, Swedish University of Agricultural Sciences, SE-75007 Uppsala, Sweden Holmboe & Skarp AB, Norr Källstavägen 9, SE-14896 Sorunda, Sweden

The authors regret that the printed version of the above article contained an error: the last row (C3) in Table 3 is incorrect. The correct

and final version of Table 3 follows. The authors would like to apologise for any inconvenience caused.

Table 3 Summary of cost and benefit values used in the cost-benefit analysis to calculate net present values. Mean PV: the mean present value of cost and benefit items. L: Large spreading scenario. S: Small spreading scenario. The annual avoided cost of inaction (B2-B3) in Alt 0-Alt 5 is assumed to be 7.5 MSEK. The social discount rate is 3.5 %. The time horizon is 120 years. Time horizon (years): 120 Discount rate: 3.5 % Category

Item

Spreading scenarios Benefit categories and items B2-B3. Avoided cost Improved health and of inaction increased provision of ecosystem services

Alt 0 S/S hotspot & disposal rest

Alt 1 S/S hotspot & S/S rest

Alt 2 S/S hotspot & stabilization AC

Alt 3 S/S hotspot & phytoremediation

Alt 4 T/T hotspot & phytoremediation

Alt 5 SW hotspot & phytoremediation

Mean PV (MSEK)

Mean PV (MSEK)

Mean PV (MSEK)

Mean PV (MSEK)

Mean PV (MSEK)

Mean PV (MSEK)

L

S

L

S

L

S

L

S

L

S

L

S

197

197

197

197

197

197

177

177

177

177

177

177

60.0

80.4

52.1

67.5

48.6

129

75.9

322

268

129

75.3

0

0

0

7.39

7.39

17.3 23.4 77.3 104 0.201 0.0346

14.9 67.0 0.0157

19.6 94.5 0.0197

10.1

9.23

6.57

Cost categories and items C1. C1a-e.I. Short-term costs 109 Remediation costs (total area) C1b.II, C1e.II. Long-term 0 costs of management, monitoring (rest of the site) C1f. Project risks 32.1 Total C1 141 C2. Impaired health C2b. From transport 0.709 due to remedial activities and C2c. At a action disposal site C3. Decreased provision of ecosystem services 82.5 due to remedial action (C3a-C3c)

79.3

7.39

7.39

7.39

13.9 16.8 69.8 154 0.0116 0.00854

13.1 74.4 70.8 16.6 96.4 404 346 153 0.00852 0.401 0.400 0

12.9 95.6 0

3.68

2.58

4.76

2.58

7.39

7.39

32.3

7.39

32.3

4.76

DOI of original article: https://doi.org/10.1016/j.scitotenv.2023.163664. * Corresponding author. E-mail address: drenning@chalmers.se (P. Drenning). https://doi.org/10.1016/j.scitotenv.2024.170429 Available online 28 January 2024 0048-9697/© 2024 The Author(s). Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).

Open Access

Advances in Environmental and Engineering Research

Review

Challenges in the PFAS Remediation of Soil and Landfill Leachate: A Review Igor Travar 1, *, Jean Noel Uwayezu 2, Jurate Kumpiene 2, Leo W.Y. Yeung 3, * 1. Ragn Sells Treatment & Detox AB, Box 952, 191 29 Sollentuna, Sweden; E-Mail: igor.travar@ragnsells.com 2. Luleå University of Technology, Geosciences and Environmental Engineering, 971 87 Luleå, Sweden; E-Mails: jean.noel.uwayezu@ltu.se; Jurate.Kumpiene@ltu.se 3. Ö rebro University, School of Science and Technology, SE-701 82 Ö rebro, Sweden; E-Mail: Leo.Yeung@oru.se * Correspondence: Igor Travar and Leo W.Y. Yeung; E-Mails: igor.travar@ragnsells.com; Leo.Yeung@oru.se Academic Editor: Zed Rengel Special Issue: Remediation of PFAS (Per- and Polyfluoroalkyl Substances) Contamination Adv Environ Eng Res 2021, volume 2, issue 2 doi:10.21926/aeer.2102006

Received: December 15, 2020 Accepted: March 19, 2021 Published: April 19, 2021

Abstract The use of per-and poly-fluoroalkyl substances (PFAS) has resulted in the contamination of different environmental matrices. In EU countries, the sites contaminated with PFAS are usually remediated by excavating the soil and disposing of it in a landfill, as no in-situ or onsite techniques capable of treating large quantities of soil cost-effectively have been developed. Landfilling of PFAS-contaminated soil is one of the sources of PFAS in landfill leachate. In this paper, the physical and chemical treatment methods to remove PFAS from soils and landfill leachates are described. Among the challenges that may limit the remediation of contaminated sites, we highlight the lack of strict regulation of PFAS in soils, the cost, the ineffectiveness of some methods for the remediation of certain PFAS compounds, and the limitation of the environmental matrices.

© 2021 by the author. This is an open access article distributed under the conditions of the Creative Commons by Attribution License, which permits unrestricted use, distribution, and reproduction in any medium or format, provided the original work is correctly cited.

Adv Environ Eng Res 2021; 2(2), doi:10.21926/aeer.2102006

Keywords Contaminated soil; landfill; leachate; treatment techniques; sorption; stabilization

1. Introduction Per-and poly-fluoroalkyl substances (PFAS) have been widely used for more than 50 years as surfactants, coatings, water repellents for leather and textiles, impregnating agents, as well, in metal plating and aqueous film-forming foams (AFFFs) [1-3]. Today, more than 4,730 compounds related to PFAS are registered, and some of them are very mobile in the environment [1, 4]. These substances have been detected in surface water [5-7], sediments downstream of the production facilities [8, 9], groundwater [10, 11], soil [12-16], wastewater treatment plant (WWTP) effluents [17, 18], and sewage sludge [19, 20]. PFAS contamination in aquatic environments has been linked to major production sources [21, 22]. Additionally, effluents from WWTPs are another major source from which PFAS are introduced into the aquatic environment [22]. In general, PFAS are fluorinated aliphatic compounds with strong carbon and fluorine (C-F) bonds. The hydrogen atoms on all the carbon atoms have been replaced with fluorine in the perfluoroalkyl compounds, whereas hydrogens on some but not all carbon atoms are replaced with fluorine in the polyfluoroalkyl compounds (which also contain a perfluoroalkyl moiety, CnF2n+1).

Perfluoroalkyl sulfonic acids (PFSAs, CnF2n+1SO3H), with 4-10 fully fluorinated carbon chains, and perfluoroalkyl carboxylic acids (PFCAs, CnF2n+1COOH), typically with 3-15 fully fluorinated carbon chains, are the most commonly reported subgroups of PFAS [23]. Long-chain PFAS with ≥8 carbon atoms are referred to as PFCAs, and those with ≥6 carbon atoms are referred to as PFSAs [24]. These longchain compounds are characterized as non-biodegradable, highly persistent, potentially toxic, and bio-accumulative (in blood, liver, and kidneys) [25]. For these reasons, together with their widespread occurrence, PFAS have been highlighted as environmental and health hazards in recent years. Two compounds, in particular, perfluorooctanoic acid (PFOA, C7F15COOH) and perfluorooctane sulfonic acid (PFOS, C8F17SO3H), have been identified as the most abundant and widespread species in soil and aquatic environments [7, 11]. Excavation and landfilling are two conventional methods for the proper management of contaminated soils, although these are becoming less popular nowadays due to their high cost.

Landfilled soils and other waste materials (e.g., carpets, textiles, paper, and packaging) containing PFAS may release these compounds in the leachate [1, 26, 27], which can then migrate to the surrounding aquatic sources, particularly groundwater, due to the lack of barriers at old landfill sites or due to damage to the bottom liners [28]. Traditional facilities for leachate treatment at the landfills in Sweden (and worldwide) are inadequate in removing PFAS [1, 3, 29]. This review aims to evaluate the available techniques for PFAS treatment in soils before landfilling and the potential release of PFAS through landfill leachate, as well as to discuss the possible challenges of implementing these techniques for the remediation of contaminated sites. We have reviewed the current legal concentrations of PFAS in landfill soil and suggest revising the legislation on the landfilling of PFAS-containing materials. The reported concentrations of PFAS in the soil in contaminated areas were compared to assess the differences among contaminated sites and discuss the need for soil treatment. The chemical stability of the treated soil in landfills is also Page 2/30

Adv Environ Eng Res 2021; 2(2), doi:10.21926/aeer.2102006

discussed and some knowledge gaps are identified. The concentrations of PFAS in landfill leachates in both new and old landfills are compiled to identify the differences among landfills globally and their environmental impacts. 2. PFAS Regulations In 2004, the EU introduced Directive 850/2004 on Persistent Organic Pollutants (POPs), which restricted the marketing and use of PFOS [30]. The Directive applies to the substances and preparations of POPs with concentrations equal to or higher than 0.005% by mass (Table 1). A summary of the applicable regulations for PFAS in soil, waste, and groundwater is provided in Table 1. The Stockholm Convention includes PFOS and PFOA in the list of POPs, and these substances are restricted under the EU POP regulations [31]. PFHxS, which is currently being assessed under REACH for potential restriction, is also considered a candidate to be listed under the POPs by the Stockholm Convention [32]. No official guideline values have been published for the PFAS concentrations in soil that could pose a potential risk to human health or the environment, either at the EU level or the national level by the member states.

The Swedish Geotechnical Institute (SGI) and the Swedish EPA have established preliminary guideline values for highly fluorinated substances in soil and groundwater, which recommend the maximum values of 0.003 mg/kg of PFOS in the soil for sensitive land use (e.g., residential land), 0.02 mg/kg of PFOS for less-sensitive land use (e.g., industrial land), and 0.045 µg/L of PFOS [33] in groundwater. In the US [34], different states have established different soil screening levels for groundwater protection which range from 0.00022 to 0.05 mg/kg of PFOS and 0.0015 to 19 mg/kg of PFOA, while the human health soil screening levels range from 0.5 to 6 mg/kg of PFOS and 0.33 to 16 mg/kg of PFOA (Table 1). In Canada, Health Canada has issued a direct contact residential soil screening value of 0.70 mg/kg for PFOA [35], and the Canadian Council of Ministers of the Environment has drafted guidelines for the protection of groundwater used for watering livestock and irrigation with the recommended PFOS limits of 12 mg/kg (coarse soil) and 9 mg/kg (fine soil) [36]. Additionally, the Canadian Federal Soil Quality Guidelines (FSQGs) have prescribed the maximum concentrations of 0.01 mg/kg of PFOS in agricultural, residential, or parkland soil, 0.14 mg/kg of PFOS in coarse soil, and 0.21 mg/kg of PFOS in fine soil at the industrial sites [37].

In Australia, several guidelines for human health soil screening criteria have been issued: 1) residential areas with garden/accessible soil: 0.009 mg/kg of PFOS and 0.1 mg/kg of PFOA; 2) residential areas with minimal opportunities for soil access: 2 mg/kg of PFOS and 20 mg/kg of PFOA; 3) public open spaces: 1 mg/kg of PFOS and 10 mg/kg of PFOA; and 4) industrial/commercial areas: 20 mg/kg of PFOS and 50 mg/kg of PFOA, which represent the values applied to PFOS as well as to the summed concentrations of PFHxS and PFOS [38]. Although several reference values are available for soil, further assessment, such as a leachability test, should be performed to evaluate the potential risk of PFAS leaching from the soil to the surrounding environment. Waste classification in the EU is regulated by the Waste Framework Directive 2008/98/EC (which categorizes waste into inert, non-hazardous, and hazardous groups) [39] and the list of waste substances in Decision 2000/532/EC [40]. The EU Regulation 850/2004 (Appendix 7) for the waste management of POPs stipulates that the waste containing more than 50 mg/kg of PFOS should be destroyed or converted irreversibly, although it does not specify where this waste should be disposed of [41]. This may cause extremely high concentrations of PFOS in the landfills, which are the sources of PFAS contamination in landfill leachate.

Moreover, inappropriate landfilling Page 3/30

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(e.g., highly contaminated waste disposed at landfills for inert or non-hazardous waste) may pose a high risk of PFAS leakage into the environment. Legal concentrations of PFAS in waste need to be revised, and maximum limits of PFAS concentrations in treated leachate should be established to minimize the risk of contamination of the recipients. Table 1 Summary of the available regulations for PFAS in soil, waste, and groundwater. Country

EU

Regulation

Subject prohibiting, phasing out or restricting the production, placing Directive 850/2004 on the market and on Persistent Organic use of Pollutants (POPs) substances subject to the Stockholm Convention on POPs

Reference

Waste containing more than 50 mgPFOS/kg TS [30, 41] should be destructed

PFOS<3 µg/kg in soil for sensitive land use Suggested values of PFOS<20 µg/kg in soil for under PFOS in soil and less sensitive land use [33] ground water PFOS<0.045 µg/L in groundwater

Sweden

Regulation construction

USA

Regulations, Guidance, and Advisories for Perand Polyfluoroalkyl Substances

Canada

Limit values

Updates to Health Canada Soil Screening Values for PFAS & Canadian Environmental Protection Act

PFOS in soil from 0.22 to levels of PFOS and 50 µg/kg PFOA in soil for PFOA in soil from 1.5 to [34] ground water 19000 µg/kg protection Levels of PFOS and PFOA in soil for human health protection

PFOS in soil from 0.5 to 6 mg/kg PFOA in soil from 0.33 to [34] 16 mg/kg PFOA

Concentrations of PFOA in residential and parkland soil, agricultural soil and soil at industrial sites

Residential, parkland and agricultural soil PFOA<0.70 mg/kg and PFOS<0.01 mg/kg Industrial sites [35, 37] PFOS<0.14 mg/kg in coarse soil and PFOS<0.21 mg/kg in fine soil Page 4/30

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Concentrations of PFOS in water used PFOS<12 mg/kg (coarse for livestock soil) watering and [36] PFOS<9 mg/kg (fine soil) irrigation to protect groundwater Residential soil: PFOS = 9 µg/kg & PFOA = 100 µg/kg; Residential with minimal Commonwealth opportunities for soil Guidelines for Environmental access: PFOS = 2 mg/kg & human health soil Australia Management PFOA = 20 mg/kg [38] screening criteria Guidance on PFOS Public open space: PFOS = and PFOA 1 mg/kg & PFOA = 10 mg/kg Industrial/Commercial: PFOS = 20 mg/kg & PFOA = 50 mg/kg 3. PFAS in Contaminated Soil Fire-fighting training areas and manufacturing plants are the largest sources of PFAS in soil. PFOS are the most intensively studied PFAS contaminants in the soil at fire-fighting training facilities [9, 16, 17]. However, the contribution of PFOS to total PFAS at the sites may be over-estimated as there would be several other PFAS that were not previously included in the analysis. These overlooked PFAS might be the novel ones, such as chlorinated, dechlorinated, and ketone PFSA, perfluoroalkane sulphonamides (FASAs), 6:2 fluorotelomer sulphonamidoalkyl betaine, 6:2 and 8:2 fluorotelomer sulfonates, or short-chain perfluorocarboxylic acids [6, 42].

In a study published in 2014 on the chemical analysis of selected fire-fighting foams from the Swedish market, PFHxA and 6:2 FTSA were found in considerably high concentrations in the foams, with some 6:2 fluorotelomer-based products also identified in most of the foam samples [43]. The reported ranges of PFOS concentrations in soil were 6.4-2400 µg/kg at Oslo airport, Norway [16], 21-120 µg/kg at Kristianstad, Bergen, and Hastad airports in Norway [16], 2.2-8520 µg/kg at Tullinge, Sweden [9], and 21-2400 µg/kg at Ellsworth Air Force Base, USA [17]. Additionally, an extremely high concentration (36000 µg/kg) of PFOS was reportedly found near a burning pit at a former firefighter training area at the Ellsworth Air Force Base [44]. The PFAS concentrations and the number of contaminated sites across the world suggest that soil is a major environmental reservoir of PFAS [15]. Hale et al. [16] and Filipovic et al. [9] found the highest concentrations of PFOS at 1-2 m below the soil, suggesting that PFOS can gradually migrate down the layers of soil with the rainwater. This migration of PFAS through the soil is complex and depends on several factors, such as the properties of PFAS, the soil properties, and the climatic conditions. Depth-profile data and mathematical modeling have demonstrated that the PFAS in the soil can be retained in the vadose zone for Page 5/30

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decades, mostly due to the adsorption of PFAS at the air-water interface as well as on soil grain surfaces [14, 15]. The transport of PFAS through the soil may result in the contamination of groundwater, surface water, and drinking water wells, along with the contamination of biota. Filipovic et al. [9] reported large differences in the concentrations of PFOS between the surface water (1.1-79 ng/L) and groundwater (1-51000 ng/L) close to an old air-force base at Tullinge near Stockholm [8]. The contamination of groundwater may occur over decades or centuries since contaminated soil acts as a reservoir of PFAS and the transport of PFAS through soil is quite slow [14, 15]. Thus, it is necessary to remediate the contaminated areas to prevent contamination of watercourses. One of the most common techniques for the remediation of the hotspots contaminated with PFAS is “digging and dumping”. Although this strategy entails costly landfilling, sometimes without proper handling and disposal facilities, its application has been continued due to the deficiencies in and the uncertainties of in-situ (on-site) techniques for treating PFAS-containing soils. Disposal of contaminated soil in a landfill from a site without treatment is merely a transfer of the problem of one site to another.

Thus, PFAS-contaminated soils need to be treated before disposal to minimize further environmental problems associated with the contamination of non-contaminated areas and drinking water sources. 4. Influence of the Landfill Environment on the Release of PFAS from Contaminated Soil The solubility of PFAS in the contaminated soil in a landfill should be kept low to avoid contamination from leachates. However, many factors in a landfill may influence the release of PFAS from the soil. Infiltration of the precipitation into the system, the type of waste, the degradation of disposed waste, and the operating conditions (e.g., compaction of wastes, gas collection system, leachate recirculation) [45] are the factors that can affect the pH, redox conditions, temperatures, and stability of PFAS in landfills [26, 46-49]. The infiltration of precipitation, moisture content, leachate production, and release of PFAS from the soil in landfills are positively correlated [26, 27]. Thus, the installation of a landfill cover reduces the production of leachate and the release of PFAS. Moreover, the reduction of infiltration reduces the microbial activity (and gas production) in the landfills, which may, in turn, reduce the release of PFAS [50, 51].

Encapsulation of PFAS-contaminated soil (e.g., with a geomembrane or clay minerals) after disposal in a landfill (during active landfilling) may prevent or reduce the contact between PFAS and the leachate, and the subsequent release from the soil. However, this technique may not have longterm (several decades to centuries) efficacy due to the potential for damage to the geomembrane and percolation of the leachate through the layer of clay [52]. Degradation of organic waste occurs in several landfilling phases, including the aerobic, acidogenic, and methanogenic phases, in which the leachate has characteristic pH values and composition (both organic and inorganic) [53]. Alkaline soils may promote the leaching of PFAS from the soil since the sorption of PFAS to the soil has been shown to decrease with an increase in the pH of the leachate. This is due to a reduction in the protonation of the adsorbent surfaces and the associated reduction in the frequencies of positively-charged sites on the sorbent [54-56]. Thus, landfilling of PFAS-contaminated soil together with alkaline materials should be avoided. The anaerobic microbial biotransformation of precursors (e.g., polyfluoroalkyl phosphate esters) may generate fluorotelomer alcohols, followed by perfluorinated/polyfluorinated acids but is Page 6/30

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probably not a major source of the PFCAs detected in anaerobic environmental matrices under methanogenic conditions [57]. This observation suggests that PFAS-contaminated soil in landfill cells, together with other biodegradable organic wastes, may reduce the degradation of PFAS precursors. However, a shift in conditions from reduction to oxidation can be expected as landfills age. Oxidation of precursors in a landfill may generate more contaminated leachate. For example, Houtz et al. [17] estimated that the oxidation of precursors accounted for 23% and 28% of total PFAS on average in groundwater and soil samples, respectively. Therefore, it is important to maintain a low redox potential in a landfill to reduce the mobility of PFAS. Landfill leachate has a complex chemical composition, and the effects of its composition on the leaching of PFAS from the waste into the landfills are not well documented. For example, anions in leachates (e.g., Cl–, SO42, Cr2O72–) and dissolved organic content (DOC) may increase the mobility of negatively-charged PFAS as they compete for the same (positively charged) adsorption sites [47, 58]. This suggests that variation in leachate quality likely contributes to the variability and patterns of PFAS concentrations in leachates [28]. Operating temperatures in a landfill may also affect the mobility of PFAS. For example, Kim et al.

reported enhanced release of PFAS (especially PFHxA and PFHpA) with an increase in the contact time and temperature [59], whereas Jia et al. found that sorption of PFOS on humic acids (the main components of organic matter in sediments) doubled when the temperature increased from 5 °C to 35 °C [49]. Thus, due to inconsistencies in the findings, more research is needed to clarify the effects of temperature on the release of PFAS from contaminated soils. Some PFAS that leach from the soil are likely to be re-adsorbed as the leachate percolates through a landfill, for example, on carbonaceous surfaces or humic acids in the waste [49]. Such readsorption has not been addressed in any study, although analogous sorption of PFAS on the sediments near WWTPs (particularly long-chain PFAS) has been reported [21, 60]. Thus, waste in a landfill could potentially serve as a filter for the PFAS leached from contaminated soils, although further research is required to test this hypothesis. 5. PFAS in Landfill Leachate Around 140-188 million tons of total annual municipal solid waste is disposed of in landfills in the USA and Europe [61, 62]. In Europe and Australia, waste containing 50 mg/kg of PFOS can be disposed of in landfills without any pre-treatment [41, 63].

In the USA, the calculated annual PFAS disposal rate is 1,250 kg/year for municipal solid waste and 470-590 kg/year for biosolids [64]. The waste containing PFAS spreads in landfill leachates. A total of 17 studies have reported the concentrations of PFAS in landfill leachates, including seven conducted in North America [27, 29,45, 51, 65-67], seven in Europe [1, 64, 68-72], two in China [3, 73], and one in Australia [74]. The concentrations of PFAS in the leachates from active landfills vary by several orders of magnitude. For example, total reported PFAS (∑PFAS) concentrations in the leachates from Canadian and Chinese landfills ranged from 27 to 21,300 ng/L and 7,280 to 290,000 ng/L, respectively [3, 29, 67, 75]. This variation is less pronounced in closed landfills than active landfills, as was reported in a study on Australian landfills (2016). Lang et al. [27] found that the concentrations of only six (PFNA, 8:2 FTCA, 5:3 FTCA, PFBS, MeFBSAA, and MeFOSAA) of the 70 PFAS analyzed were significantly higher in the younger US landfills than in the older ones. This could be due to reductions in the concentrations of these species with time or changes in the types of PFAS used in products in recent Page 7/30

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years (e.g., a shift to the use of PFBS-based and fluorotelomer-based products instead of PFOS and PFOA). The similarity in the concentrations of almost all the analyzed PFAS in the leachates from old and young landfills indicates that PFAS may be released for many years after the disposal of the soil. Concentrations of PFAS in landfill leachates and volumes of accumulated leachates must be known when evaluating the hazards posed by landfills, particularly the risk of PFAS contamination to the aquatic environment. For example, Busch et al. [1] estimated that discharges of ∑PFAS from 22 landfills in Germany ranged between 0.08 and 956 mg/day (mean: 135 mg/day), and the total discharges from all landfill sites in Germany (ca. 1,700) amounted to approximately 240 g of total PFAS/day, on average. Fuertes et al. [71] estimated a 1,209 g/year mass flow of ∑PFAS from four studied landfill sites serving a population of about 1.8 million people in northern Spain. In the USA, landfills released 563-638 kg of ∑PFAS to WWTPs via leachates in 2013 [27]. The actual amount of PFAS in landfill leachate in the USA could be higher as the cited authors’ analyses did not include total oxidizable precursor or total fluorine assays. Furthermore, Yan et al. [3] estimated that approximately 3,110 kg of PFAS is released in landfill leachates in China annually.

Thus, landfill leachates can be a significant source of PFAS contamination in the environment. PFAS should, thus, be removed from leachates before discharging to recipients, such as WWTPs, rivers, or lakes [65]. 6. Treatment of PFAS-Contaminated Soil before Landfilling 6.1 Removal of PFAS from Contaminated Soil through Washing Soil washing involves the separation of PFAS-contaminated particles from the bulk of the soil, followed by the extraction or desorption of PFAS from those particles (Figure 1). Mechanical separation is suitable for coarse-grained soils (e.g., soils dominated by stones and gravel), but it is not economically viable for fine-grained soils (dominated by fine sand, silt, and clay) [16]. Chemical extraction of PFAS should be performed on fine-grained soil particles under saturated conditions to enhance the treatment and minimize the adsorption of PFAS in unsaturated soils (see Section 3). Extraction agents, such as methanol/sodium hydroxide (MeOH/NaOH), methanol/ammonium hydroxide (MeOH/NH4OH) and acetonitrile (CH3CN), and water, have been used to extract PFAS from contaminated soils in the laboratory under saturated conditions [16, 42, 54]. Hale et al.

[16] found that in the determination of the total concentration of PFOS in the soil, water extracted more PFOS compared to either methanol or acetonitrile, indicating that water could serve as an excellent extraction agent. However, extraction efficiency depends not only on the type of extraction agent but also on the soil properties (e.g., organic matter and clay content) [16, 60]. Removing PFAS by washing may be feasible for soils with low DOC (0.7-6 mg/L) and soils with low clay content (e.g., soils predominantly composed of sand) at an alkaline pH [16, 60]. In addition to the treated soil, this technique generates large amounts of contaminated process water that needs further treatment. Soil washing has been used to remove PFAS commercially in-situ (“flushing”) but not for the treatment of soil before landfilling due to the high costs involved [54].

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Concentrated PFASs in soil

PFAS-contaminated soil

Mechanical separation

Clean soil particles

Process water/Extracting agent Extraction of PFASs with extraction agent

Process water contaminated with PFASs

Treated/clean soil

Treatment of process water PFAS contaminants

Figure 1 A combined treatment of PFAS-contaminated soil consisting of mechanical separation of the contaminated particles from the soil mass and the extraction of PFAS using an appropriate agent. 6.2 Stabilization and Solidification of PFAS in Contaminated Soil Stabilization of PFAS in contaminated soil refers to the immobilization of PFAS by the addition of a substance that reduces mobility through leaching. Solidification of PFAS-contaminated soil involves the mixing of a cementitious binder and additives into the contaminated matrix to reduce hydraulic conductivity and PFAS exposure to leaching. Detailed knowledge of PFAS sorption behavior is essential for developing effective stabilization techniques to manage contaminated soils. The most important types of interactions involved in the sorption of PFAS to potential sorbents are electrostatic interactions, hydrophobic interactions, and ion exchange [56, 76, 77]. Electrostatic interactions between PFAS in solution and sorbents rely significantly on the pH [46-48, 78]. Anionic PFAS are adsorbed on the positively-charged surfaces of the sorbents [47-49, 79]. When the pH of a PFAS-contaminated solution becomes more acidic than the adsorbent’s pHpzc (pH at the point of zero charge), the number of positively-charged sites on the sorbent increases (by surface protonation), and the adsorption of these PFAS is enhanced [76].

Hydrophobic interaction between PFAS and sorbents is the strongest when the sorbents are uncharged [60, 76]. Ion exchange is the dominant mechanism for the immobilization of anionic PFAS, especially for short-chain PFAS, when the anion-exchange resin is used [73]. The described sorption mechanisms of PFAS on sorbents are quite complex as the immobilization of PFAS in stabilized soil depends not only on the properties of the added sorbent but also on the soil characteristics (e.g., organic matter content) and leaching conditions (e.g., pH, redox potential, and temperature). Both inorganic and organic substances have been tested as possible PFAS adsorbents under experimental conditions, mostly in soil-free conditions. The inorganic substances include sand, high iron sand, kaolinite, Fe 2O3 (hematite), alumina, zeolite, hydrotalcite, montmorillonite, FeO(OH) (goethite), and modified commercial clay [48, 73, 80, 81]. The organic substances include oil, black carbon, activated carbon (AC), biochar (BC), and ion-exchange resin [16, 82-85]. However, only four of the reviewed studies [16, 80, 84, 86] included tests evaluating the PFAS sorption efficiency in the presence of soil and only one published study described the solidification of PFAS-contaminated soil [86].

The stabilization efficiency of the substances should be verified in the presence of soil containing multiple contaminants collected from a field site.

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A

B

C

Composition % Concenteration ng/L Concenteration ng/L

Adv Environ Eng Res 2021; 2(2), doi:10.21926/aeer.2102006

1000000

100000 10000 1000 100 10

1 35000

350000 29080

30000

288559 26024

300000

25000

250000

20000

200000

15000 13920

13346

150000

10000

7600

5000

3720

1731.3 613.7

0

146.1

100000 50000

2285

0

100%

PFBS

90%

PFHxS

80%

Singapore (n=12)

Spain (n=4)

PFDA

Germany (n=6)

PFNA

0%

Finland (n=2)

PFOA

10%

Norway, Fin. (n =9)

PFHpA

20%

Sweden (n=29)

PFHxA

30%

China (n=5)

PFPeA

40%

Aust. (n=32) close land. Aust. (n=28) oper. land.

PFBA

50%

US (n=12)

PFOS

60%

Canada (n=68)

70%

Figure 2 Graphs showing (A) Graphs showing minimum and maximum concentrations (ng/L) of ∑PFAS in raw leachate; (B) Graphs showing average concentrations (ng/L) of ∑PFAS in raw leachate. (The bar in black does not fit onin the y-axis.); and (C) Graphs showing the composition of PFAS (%) in the raw leachate. Ref. From earlier studies: Canada [29, 67, 75], the USA [45, 51], Australia [74], China [3], Sweden [64, 72], Norway and Finland [68], Finland [70], Germany [1], Spain [71], and Singapore [87]. 6.2.1 Treatment of PFAS-Contaminated Soil Using Activated Carbon (AC) and Biochar (BC) Biochar is a type of charcoal produced mainly from biomass (e.g., mixed wood, coconut husks, or bamboo) through pyrolysis (thermal decomposition of organic matter in the absence of oxygen). Granulated AC (GAC) and powdered AC (PAC) are produced similarly (using biomass, coal, and peat), in combination with physical activation (using hot gases) or chemical activation (with acid or base, such as H3PO4, KOH, and NaOH), to increase the pore volume and surface area. Both AC and BC treatment technologies have been used in the remediation of soil, groundwater, and drinking water contaminated with PFAS [16, 83, 84, 88]. AC is more suitable for remediation as its pore structure provides considerably large specific surface areas, pore volumes, and sorption capacities for organic pollutants (Table 2).

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Table 2 Chemical composition, surface area (SA), pore volume (PV), soil-water distribution coefficient Kd (in L/kg), and organic carbon normalized sediment-water distribution coefficient logKOC (estimated as Kd/foc) of soil, and the soil treated with powdered activated carbon (PAC), mixed wood biochar (MW BC), paper mill waste biochar (PMW BC) [8, 16, 76, 84, 89]. Soil/sediment

PAC 9.6 [84]

MW BC 9.6 [84]

pH Chemical composition, % C 81 [84] 53 [84] H 0.03 [84] 0.8 [84] N 0.2 [84] 0.3 [84] O 8.4 [84] 2.5 [84] Physical properties SA, m2/g 346-812[16, 76, 84] 404-520 [84] PV, cm3/g 0.27-0.6 [84] 0.14-0.19 [84] Soil-water distribution coefficient logKd (L/kg) PFBA (C4) 1.1-<2.3 [84] 3.01 [84] 1.1-<2.2 [84] PFHxS (C6) 1.2-1.8 [84] 3.5 [84] 1.4-1.8 [84] PFOS (C8) 1.3-3.8[8, 16, 76, 89] 3.8-4.6[16, 76, 84] 1.8-3.1 [84] PFOA (C8) 0.3-4 [76, 89] 3.3-4 [76, 89] 2.2-4 [84] PFNA (C9) 2.3-<2.6 [84] >3 [84] 2.2-<2.8 [84] Organic carbon normalized sediment-water distribution coefficient logKOC (L/kg) PFBA (C4) 1.8-<4.8 [84] 4 [84] 2.4-4.8 [84] PFHxS (C6) 1.8-4 [84] 3.9-5.8 [84] 1.8-4[84] PFOS (C8) 2.9-4.3 [76, 89] 4.3-7.2 [76, 89] 2.9-4.5[80] PFOA (C8) 1.9-2.6 [8, 76, 89] 4.4-5.5 [76, 89] 2-4.7 [84] PFNA (C9) <2.8-<5.3[84] >4 [84] <2.9-<5 [84]

PMW BC 9.5 [84] 19 [84] 0.5 [84] 0.1 [84] 9.4 [84] 13.2-48 [84] 0.01-0.09 [84] 1-<2.2 [84] 1.3-1.8 [84] 1.8-3 [84] 3-4 [84] 2.1-<2.8 [84] 2.3-4.8 [84] 1.8-4 [84] 3.1-4.4 [84] 2.1-4.6 [84] <3.1-<5.5 [84]

Hale et al. [16] observed that the addition of 3% PAC (in a specific area of 726 m2/g) to contaminated soil in batch leaching tests reduced the leaching of PFOS by 94%-99.9%, from 6.4-54.5 µg/kg to 0.003-0.43 µg/kg. Modeling of these results using a one-dimensional box model indicated that this would reduce the concentrations of PFOS in pore water from 19.5 µg/L to 0.4 µg/L, just above the clean-up target of 0.3 µg/L. The same model indicated that PFOS would be released slowly from the untreated soils, taking more than 100 years to decrease the pore water concentrations of PFOS below the arbitrary target. This suggests that PAC might serve as an effective sorbent to diminish releases from contaminated soil to the surrounding environment, although more information is needed to confirm its efficacy. For example, further investigation is needed to determine the suitability of the method for field application, PAC’s durability in soil, and the subsequent release of immobilized PFAS (i.e., the long-term treatment efficiency), as well as, the material costs and the wider environmental impacts of the process. Biochar could immobilize PFAS in solution in sorption experiments (as shown by the Freundlich sorption coefficients presented in Table 2) but could not immobilize PFAS in soils in batch leaching Page 11/30

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tests [84]. This is probably due to the blockage of biochar pores with dissolved compounds and particles (e.g., humic substances and oil) in the soil [84]. Moreover, other soil components (e.g., DOC and ions) can compete with PFAS for the sorption sites [47, 90, 91]. The effectiveness of BC in immobilizing PFAS may be higher in the soils with lower organic content (e.g., total organic carbon (TOC)<3% in dry mass), indicating that for successful amendment, the amount of carbonaceous material added to the contaminated soil should exceed the initial carbon content of the soil [84, 92]. Thus, BC can be used for the sorption of PFAS in soils with low organic content. The effectiveness of soil treatment with carbonaceous amendments (CAs), such as PAC and BC, depends not only on the CA but also on the chain-length of the PFAS. For example, the logarithm of the distribution coefficient (logKd) of PFBA in the soil stabilized with PAC is about 3.01 L/kg, whereas, the logKd of PFOS in the same system varies between 3.8 and 4.8 L/kg [13, 28] (Table 3). These results show that PFAS with longer chains (e.g., PFOS and FOSA) bind more strongly to CA than those with shorter chains, indicating the importance of hydrophobic interactions in the sorption of PFAS (Higgins and Luthy, 2006).

Moreover, PFAS with a sulfonate moiety have stronger adsorption than those with a carboxylate moiety (e.g., in the soil stabilized with PAC, the logK d of PFOS reportedly varies between 3.8 and 4.6 L/kg, while the logKd of PFNA is about 3; see Table 2). This is probably due to the slightly higher hydrophobicity of the slightly larger sulfonate moiety and/or differences in the specific electrostatic interactions of these moieties [60, 84]. Moreover, short-chain PFAS and fluorotelomers (which have a partially fluorinated carbon chain) can be more easily desorbed from the sediments with a low fraction of organic carbon (e.g., f oc of 0.03%) in the aqueous phase than the species with longer chains or fully-fluorinated carbon chains (e.g., PFOS and FOSA) [84]. For example, Ahrens et al. [8] and Kupryianchyk et al. [84] calculated the carbon-normalized distribution coefficients and log (Kd/foc) values of 2.9-4.3 L/kg for PFOS and 2.1-2.6 L/kg for PFOA. Thus, it is important to add enough CA to PFAS-soil to ensure the stability of immobilized substances. Normally, PFAS-contaminated soils are remediated with 2%-4% CA (about 30-60 t/ha, assuming mixing to a depth of ca. 10 cm and soil density of ca. 1.5 kg/m3) in-situ [11]. Whether this amount of CA is sufficient to stabilize soils in landfills remains unknown as no relevant information was found in the reviewed papers.

Table 3 Comparison of ∑PFAS in untreated and treated leachate with respective treatments [1, 87]. WAO-wet air oxidation, AC-activated carbon, CW-constructed wetlands. Treatment method WAO AC AC AC CW

∑PFAS ∑PFAS in treated untreated leachate, ng/L ng/L 1889 1993 31 9 12819 4079 8370 20 814-4324 367-954

in leachate, Landfill status inactive inactive inactive active active

Reference [1] [1] [1] [1] [87]

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6.2.2 Treatment of PFAS-Contaminated Soil with Compost Hale et al. [16] found that treating contaminated soil with 3% compost (composed of peat and cow manure with a high content of nitrogen, phosphate, and potassium) reduced PFOS leaching by 29%-34%, from 3.8 ± 2.4 to 2.54 ± 1.5 µg/L; the calculated distribution coefficient (Kd) of PFOS between compost and soil was 8.8 L/kg. Sorption of PFAS on compost occurs probably through direct polar interactions, hydrophobic interactions, ion-exchange, and addition to more specific sites of moieties with an affinity for PFAS present in the organic matter [16, 79]. More information is, nonetheless, needed on the sorption of PFAS using compost, the influence of co-contaminants, the stability of immobilized PFAS, and the efficacy of the treatment. 6.2.3 Treatment of PFAS-Contaminated Soil with Clay Minerals (Montmorillonite and Kaolinite) Soil rich in clay minerals, such as montmorillonite and kaolinite, can be used to stabilize PFAScontaminated soil. For example, Hale et al. [16] found that treating contaminated soil with 3% montmorillonite decreased the release of PFOS by 28%-40%, from 3.8 ± 2.4 µg/L to 2.7 ± 1.8 µg/L. The reported Kd values for the sorption of PFOS to montmorillonite were in the range of 5–10 L/kg [16, 93].

Hydrophobic and electrostatic interactions are the main mechanisms involved in the adsorption of PFOS to montmorillonite and kaolinite. More specifically, PFOS may be immobilized via the formation of outer-sphere surface complexes and specific interactions of the sulfonate group of PFOS with the hydroxyl groups on the surfaces of clay minerals [94]. However, dissolved humic substances in soil were shown to hinder the adsorption of PFOS on clay minerals via electrostatic repulsion and occupation of adsorption sites [93]. Thus, treatment with clay minerals can be effective for PFAS-contaminated soils with low organic matter, such as soils predominantly composed of stones and sand. Moreover, the addition of clay particles to such soil probably reduces its hydraulic conductivity, and hence, the contact between the contaminated soil and infiltrating water. Further research is needed on the sorption of PFAS in the presence of soil, the influence of co-contaminants, and the stability of immobilized PFAS. 6.2.4 Solidification of PFAS-Contaminated Soil Sörengård et al. [86] tested the solidification of PFAS-contaminated soil with a combination of Portland cement (PC), fly ash (FA), and ground granulated blast furnace slag (GGBS) with a PC:FA:GGBS ratio of 1:1:2.

This solidification was done together with chemical stabilization using several additives, such as PAC, Rembind® (AC-based and containing amorphous aluminum hydroxide and kaolin clay), pulverized zeolite, chitosan, hydrocalcite, and bentonite at a concentration of 0.2% (w/w) of dry PFAS-contaminated soil. The results showed that most PFAS (except the PFCAs with a perfluorocarbon chain length of less than five carbon atoms, e.g., PFBA and PFPeA) were successfully stabilized with CA in the solidified soil (e.g., logK d increased from 1.5 L/kg for PFOS in the soil to about 3.5 L/kg treated with PAC) by hydrophobic sorption. Sörengård et al. recently demonstrated the long-term stabilization of PFAS in AFFF-contaminated soil, with a removal rate of up to 97% for the most important PFAS (PFOA and PFOS) [95]. The short-chains of perfluorocarbon could not be immobilized in significant amounts in stabilized/solidified soil (S/S-soil) under basic pH as they have a lower hydrophobic bond strength and faster diffusion in the water phase compared to the long-chains [86]. Hence, the S/S-soil, with Page 13/30

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large numbers of short perfluoroalkyl chains, needs pH adjustment to reduce leaching. Apart from the positive effects on the immobilization of PFAS, the addition of CA to soil exerts negative effects on solidification. For example, Sörengård et al. [86, 95] reported that unconfined compressive strength (as a measure of solidification) decreased by 38% and 14% in a carbon-poor sandy silt soil treated with 0.2% Rembind® and PAC, respectively. This means that the solidification of soils rich in organic carbon and clay through treatment with CA may entail high costs as these soils need significantly more binder in combination with PC and the addition of sodium silicate and/or lime compared to sandy soils [86]. Further laboratory studies should be performed before S/S-treatment is implemented in the field to optimize the S/S-binder methods for the specific contaminated soil, due to the complexity of the binding of soil to aggregates. The stability of PFAS in S/S-soil over time is not quite clear. The carbonation of FA and PC may result in a decrease in the pH and improvement in the electrostatic interactions between PFAS and the treated soil. Contrary to the positive effects of carbonation, the aging of CA in S/S-soil may result in low sorption capacity for PFAS or even in the desorption of PFAS [58].