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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 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.
 
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 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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==Advantages and Disadvantages==
 
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There are many advantages to SCWO treatment. SCWO is a destructive treatment in that the compounds treated are mineralized to simple elements or harmless molecules (e.g., water and carbon dioxide) rather than just being transferred to another medium. Another advantage is the absence of reaction by-products, incompletely oxidized contaminants or unreacted harmful oxidants (e.g., ozone). SCWO is an extremely rapid and effective reaction (typical reaction times are in the order of 5-10 seconds) making it possible to build systems that are very compact and have a high throughput. SCWO is also a very clean process. The highly oxidizing environment makes it possible to effectively treat all sorts of organic contaminants, often recalcitrant to other processes, with very high (>99%) destruction efficiencies. This includes treatment of trace contaminants, slurries of biosolids, waste oil, food wastes, plastics, or emerging contaminants such as PFAS or 1,4-dioxane. Also, the relatively moderate temperatures (380-600&deg;C) compared to other destructive technologies such as incineration, combined with the presence of supercritical water prevent the formation of NOx and SOx compounds. Lastly, SCWO treatment does not require drying of the waste, and both liquids and slurries can be treated using SCWO.
 
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There are several disadvantages to SCWO treatment. First, a significant amount of energy needs to be expended to bring the oxidant and the waste undergoing treatment to the critical point of water. Although a large fraction of this energy can be efficiently recovered in heat exchangers, compensating for heat losses constrains SCWO to the treatment of concentrated wastes with sufficient organic content for the exothermic oxidation reaction to provide the necessary heat. Typically, a minimum calorific content of around 2 MJ/kg (which generally corresponds to a chemical oxygen demand of about 120-150 g/L) is needed for autothermal operation. For more dilute streams, external heating or supplementation of fuel (diesel, alcohol, waste oil, etc.) can be implemented, but it can rapidly become cost prohibitive. Thus, SCWO is currently not economical for very large volumes (>50,000 gallon/day) of very dilute waste streams. A second limitation is related to the pumping of the waste. Because the process is conducted at high pressure (240 bars or 3500 psi), positive displacement pumps are required. This limits SCWO to liquids and slurries that can be pumped. Waste streams that contain excessive grit or abrasive materials, and soils cannot currently be processed using SCWO.
 
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The many appealing benefits of supercritical water processing have stimulated engineers and entrepreneurs to invest significant efforts and resources in the development of the technology. Today, after roughly 30 years of development, commercial deployment is on the horizon<ref name="Marrone2013"/>. Technical challenges that have slowed down commercial deployment of SCWO are linked to the complex nature of a high-pressure, high-temperature process. Critical issues include reactor materials selection to resist corrosion (typically high nickel alloys are used), reactor designs and construction to withstand the corrosive nature of the reactive mass, dealing with highly exothermic reactions at high pressure and high temperature, plugging of the reactor by minerals deposits, and energy recovery for autothermal operation. Another challenge was the unrealistic goal of some companies entering the SCWO market to produce power from waste streams (often wastewater sludge) at a competitive cost (3-5 cents/kWh). This was not feasible with the available technology, which led to several business failures.
 
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The value proposition of treating recalcitrant wastes using SCWO is markedly different, especially in today’s context of increasing liability for trace levels of emerging contaminants such as PFAS. SCWO may prove to be the optimal treatment technology for many highly concentrated aqueous waste streams.
 
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[[File: Nagar1w2Fig2.png | thumb | 500px | Figure 2.  Duke SCWO pilot-scale system (during construction, thermal insulation removed). The system is housed in a standard 20 ft shipping container and can treat about 1 ton (or 270 gallons) of waste per day.]]
 
  
 
==State of the Art==
 
==State of the Art==

Revision as of 19:50, 18 March 2021

Mercury in Sediments

Mercury (Hg) is released into the environment typically in the inorganic form. 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). Direct exposure to Hg(II) and Hg(0) can be a human health risk at heavily contaminated sites. However, the organic form of Hg, methylmercury (MeHg), is a neurotoxin that can bioaccumulate and is the form of Hg that poses the greatest human and ecological health risk. As a chemical element, Hg cannot be destroyed, so the goal of Hg-remediation is immobilization and prevention of food web bioaccumulation.

Related Article(s):

Contributor(s): Dr. Grace Schwartz

Key Resource(s):

  • Challenges and opportunities for managing aquatic mercury pollution in altered landscapes[1]
  • The assessment and remediation of mercury contaminated sites: A review of current approaches[2]
  • Bioaccumulation and Biomagnification of Mercury through Food Webs[3]

Introduction

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 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[2].

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[2]. Direct exposure to Hg(II) and Hg(0) can be a human health risk at heavily contaminated sites. However, the organic form of Hg, methylmercury (MeHg) 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 Minamata Bay, Japan (see also 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 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.

Biogeochemistry/Mobility of Hg in soils

In the environment, Hg mobility is largely controlled by chelation with various ligands or adsorption to particles[1]. 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)[1]. MeHg has mobility tendencies similar to Hg, with DOM and sulfur ligands competing with each other to form complexes with MeHg[4]. However, unlike Hg-S complexes, MeHg-S does not have limited solubility.

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- methylating microorganisms, (2) the activity of Hg-methylating microorganisms, and (3) the rate of biotic and abiotic demethylation. MeHg is produced by anaerobic microorganisms that contain the hgcAB gene[5]. 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[6]. 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)[6]. 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[7][8]. Methylmercury can be degraded biotically by aerobic bacteria containing the mercury detoxification, mer operon and through oxidative demethylation by anaerobic microorganisms[7].

Bioaccumulation and Toxicology

Regulatory criteria are most often based on total Hg concentrations, however, MeHg is the form of Hg that can bioaccumulate in wildlife and is the greatest human and ecological health risk[9]. MeHg represents over 95% of the Hg found in fish[10]. 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[11]. 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[12]. 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[11].

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[13]. 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[14].

Regulatory Framework for Mercury

In the United States, mercury is regulated by several different 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[15].

In 2013, the United States signed the international 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[2]. 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.[2]and Wang et al.[16] give thorough reviews of standard and emerging technologies.

Recently application of in situ sorbents has garnered interest as a remediation solution for Hg[2]. Many different materials, including biochar and various formulations of activated carbon, are successful in lowering porewater concentrations of Hg and MeHg in contaminated sediments[17]. 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 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.

State of the Art

Figure 3. Landfill leachate (left) and SCWO treated effluent (right). Effluent is odorless.

Relatively few large scale SCWO systems exist. Researchers at Duke University (Deshusses lab) have designed and built a prototype pilot-scale SCWO system housed in a standard 20-foot shipping container (Figure 2). This project was funded by the Reinvent the Toilet program of the Bill and Melinda Gates Foundation. The pilot system is a continuous process designed to treat 1 ton of sludge per day at 10-20% dry solids content. The unit has been undergoing testing at Duke since early 2015. The design includes moderate preheating of the waste slurry, followed by mixing with supercritical water (~600°C) and air, which serves as the oxidant. This internal mixing rapidly brings the waste undergoing treatment to supercritical conditions thereby minimizing corrosion and the risks of waste charring and associated reactor plugging. The organics in the sludge are rapidly oxidized to CO2, while the heat of oxidation is recovered to heat the influent waste. The reactor is a single tubular reactor. The high supercritical fluid velocity in the system helps with controlling mineral salts deposition in the reactor. The system is well instrumented, and operation is controlled using a supervisory control and data acquisition (SCADA) system with historian software for trends analysis and reporting of key performance indicators (e.g., temperatures and pressures, pollutant destruction). Experiments conducted with this pilot plant have shown effective treatment of a wide variety of otherwise problematic wastes such as primary, secondary and digested sludge slurries, landfill leachate (see Figure 3), animal waste, and co-contaminants including waste oil, food wastes, and plastics. The results are consistent with other SCWO studies and show very rapid treatment of all wastes with near complete conversion (often >99.9%) of organics to CO2. Total nitrogen and phosphorous removal are generally over 95% and 98%, respectively. Emerging contaminants such as pharmaceuticals, PFAS, 1,4-dioxane and microplastics are also treated with destruction generally exceeding 99%.

Early projections for treatment costs (Capital Expenditures + Operating Expenditures) for slurries are in the range of $12 to $90 per ton (or $0.04 to $0.37 per gallon) depending on system scale and contaminant concentration, with a majority of the cost coming from amortizing the equipment. These cost projections make SCWO treatment very competitive compared to other treatment technologies for high-strength wastes. When treating large volumes of water, combining SCWO with another technology (e.g., nanofiltration, reverse osmosis, or adsorption onto GAC) should be considered so that only the concentrated brines or spent sorbent are treated using SCWO, thereby increasing the cost effectiveness of the overall treatment.

Table 2. Results for influent biosolids and treated effluent using Duke University SCWO pilot-scale plant
Substance Residual
(ng/L) 		Removal
PFBA 10.20 99.86%
PFHxA 5.15 99.89%
PFNA 1.07 99.90%
PFDA 0.80 99.97%
PFUnA <1.10 >99.89%
PFBS <0.19 >99.98%
PFPes <0.29 >99.98%
PFHxS 0.28 99.99%
PFOS 0.65 99.99%
Note: Similar destruction efficiencies were obtained when treating AFFF solutions.

SCWO for the Treatment of PFAS and AFFF

Several reports have indicated that PFAS can be treated using SCWO[18]. Several runs treating biosolids known to contain PFAS as well as dilutions of pure aqueous film forming foam (AFFF) have also been conducted with the Duke SCWO system. Typical results are shown in Table 2. They indicate very effective treatment performance, with for example 110,000 ng/L PFOS in the feed reduced to 0.79 ng/L in the effluent, and many other PFAS reduced to below their detection limits. No HF was found in the effluent gas, and all the fluorine from the destroyed PFAS was accounted for as fluoride in the effluent water. These results show the ability of the SCWO process to destroy PFAS to levels well below the EPA health advisory levels of 70 ng/L for PFOS and PFOA. The Environmental Security Technology Certification Program (ESTCP) project number ER20-5350[19] launched in June 2020 will assess the technical feasibility of using supercritical water oxidation (SCWO) for the complete destruction of PFAS in a variety of relevant waste streams and will evaluate the cost effectiveness of the treatment.

References

  1. ^ 1.0 1.1 1.2 Hsu-Kim, H., Eckley, C.S., Achá, D., Feng, X., Gilmour, C.C., Jonsson, S., Mitchell, C.P.J., 2018. Challenges and opportunities for managing aquatic mercury pollution in altered landscapes. Ambio, 47, pp. 141-169. DOI: 10.1007/s13280-017-1006-7   Free access article   Report.pdf
  2. ^ 2.0 2.1 2.2 2.3 2.4 2.5 Eckley, C.S., Gilmour, C.C., Janssen, S., Luxton, T.P., Randall, P.M., Whalin, L., Austin, C., 2020. The assessment and remediation of mercury contaminated sites: A review of current approaches. Science of the Total Environment, 707, Article 136031. DOI: 10.1016/j.scitotenv.2019.136031   Free download from: ResearchGate
  3. ^ 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. DOI: 10.1002/9781118146644.ch14
  4. ^ 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. DOI: 10.3184/095422907X255947   Free access article   [Media: Loux2007.pdf | Report.pdf]]
  5. ^ 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. DOI: 10.1126/science.1230667
  6. ^ 6.0 6.1 Bravo, A.G., Cosio, C., 2020. Biotic formation of methylmercury: A bio–physico–chemical conundrum. Limnology and Oceanography, 65(5), pp. 1010-1027. DOI: 10.1002/lno.11366   Free Access Article   Report.pdf
  7. ^ 7.0 7.1 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. DOI: 10.1007/s00128-018-2530-2
  8. ^ Jonsson, S., Mazrui, N.M., Mason, R.P., 2016. Dimethylmercury Formation Mediated by Inorganic and Organic Reduced Sulfur Surfaces. Scientific Reports, 6, Article 27958. DOI: 10.1038/srep27958   Free access article   Report.pdf
  9. ^ Agency for Toxic Substances and Disease Registry (ATSDR), 1999. Toxicological Profile for Mercury. Free download   Report.pdf
  10. ^ 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. DOI: 10.1139/f92-113
  11. ^ 11.0 11.1 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. DOI: 10.1002/9781118146644.ch14
  12. ^ 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: ResearchGate
  13. ^ 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. DOI: 10.1056/NEJMra022471
  14. ^ 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. DOI: 10.1016/j.envres.2017.08.051
  15. ^ US EPA, 2021. Environmental Laws that Apply to Mercury. US EPA Website
  16. ^ 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. DOI: 10.1016/j.envint.2019.105281   Free access article
  17. ^ 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. DOI: 10.1021/es4021074   Free download from: ResearchGate
  18. ^ Kucharzyk, K.H., Darlington, R., Benotti, M., Deeb, R. and Hawley, E., 2017. Novel treatment technologies for PFAS compounds: A critical review. Journal of Environmental Management, 204(2), pp. 757-764. DOI: 10.1016/j.jenvman.2017.08.016   Manuscript available from: ResearchGate
  19. ^ Deshusses, M.A., 2020. Supercritical Water Oxidation (SCWO) for Complete PFAS Destruction. Environmental Security Technology Certification Program (ESTCP) Project number ER20-5350. Project website

See Also

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