Evaluation of a customized reactive nanoscale-zero-valent iron and zeolite thin capping blend for enhancing natural recovery of wetlands impacted by contaminated legacy gold mine tailings

Evaluation of a customized reactive nanoscale-zero-valent iron and zeolite thin capping blend for enhancing natural recovery of wetlands impacted by contaminated legacy gold mine tailings | Research Square window.SnipcartSettings = { analytics: { enabled: false } }; (function() { var accessVector = localStorage.getItem('access_vector') || ''; window.dataLayer = window.dataLayer || []; if (accessVector) { window.dataLayer.push({ user: { profile: { profileInfo: { snid: accessVector } } } }); } })(); (function(w,d,s,l,i){w[l]=w[l]||[];w[l].push({'gtm.start':new Date().getTime(),event:'gtm.js'});var f=d.getElementsByTagName(s)[0],j=d.createElement(s),dl=l!='dataLayer'?'&l='+l:'';j.async=true;j.src='https://www.googletagmanager.com/gtm.js?id='+i+dl;f.parentNode.insertBefore(j,f);})(window,document,'script','dataLayer','GTM-K279D39R'); Browse Preprints In Review Journals COVID-19 Preprints AJE Video Bytes Research Tools Research Promotion AJE Professional Editing AJE Rubriq About Preprint Platform In Review Editorial Policies Our Team Advisory Board Help Center Sign In Submit a Preprint Cite Share Download PDF Research Article Evaluation of a customized reactive nanoscale-zero-valent iron and zeolite thin capping blend for enhancing natural recovery of wetlands impacted by contaminated legacy gold mine tailings Ellen Emily Vanessa Chapman, Linda M. Campbell This is a preprint; it has not been peer reviewed by a journal. https://doi.org/ 10.21203/rs.3.rs-3894488/v1 This work is licensed under a CC BY 4.0 License Status: Posted Version 1 posted You are reading this latest preprint version Abstract Legacy gold mine tailings from the 1800’s in Nova Scotia, Canada have elevated mercury (Hg) and arsenic (As) concentrations. Tailings, were slurried into wetlands without treatment. Over a century later, those impacted wetlands are still at risk and innovative in-situ treatment approaches to support natural biological and chemical recovery are needed. Here we report results of our proof-of-concept laboratory study with a customized reactive thin layer capping to limit mobility, bioaccumulation and toxicity of Hg and As in wetland sediment impacted by legacy tailings. The customized reactive amendment is a blend of NANOFER STAR nanoscale zero valent iron (nZVI) and fine-grained zeolite (clinoptilolite) inserted either below, or within a thin cap (silica sand, bentonite and zeolite) and placed over contaminated wetland sediments in beakers. Due to the high concentrations of Hg and As in sediments, invertebrates ( Hyalella azteca , Daphnia magna and Caridina multidente) exposed to untreated wetland sediment exhibited high mortality and bioaccumulation of Hg. The reactive capping applications improved the survival of H. azteca and D. magna similar to the survival rates seen in our clean control sediment. Bioaccumulation of Hg was also reduced in C. multidente exposed to the treated sediment compared to the untreated sediment. Furthermore, total [Hg] and [As] in the overlaying water of treated contaminated sediments were reduced by 88% and 99% respectively. Our proof-of-concept testing of this reactive capping blend shows potential for managing and supporting natural recovery of wetlands impacted by historical gold-mine tailings. Toxicity mercury arsenic sediment in-situ risk management Figures Figure 1 Figure 2 Figure 3 Figure 4 Figure 5 Figure 6 Figure 7 Figure 8 Introduction With 360 gold mines in 64 historic gold mining districts across Nova Scotia (NS) Canada, there is a significant legacy of potentially toxic mining waste “tailings” in this province. It has been estimated that over 3 million tonnes of tailings were generated during the historical gold rushes in NS between the 1860’s & 1940’s (Parsons et al. 2012 ). Tailings contain elevated concentrations of mercury (Hg) due to the mercury amalgamation process used to extract the gold, and arsenic (As) due to the geogenic arsenopyrite in gold-bearing rock (Meunier et al., 2010). Gold amalgamation processes in the late 1800’s required freshwater to process crushed ore, so legacy gold mine ore processing and tailing sites are generally situated close to low-lying lakes, rivers, streams or wetlands. After ore processing, the tailing material was often slurried back into the same freshwater bodies without treatment. Consequently, aquatic ecosystems in historical gold mine districts are at particular risk, even over 100 years later. Despite this, impacts of tailings on aquatic ecosystems and freshwater biota from legacy gold mines remain poorly understood ( LeBlanc et al. 2019 ). Multi-generational chronic exposure to toxic tailings has occurred in many wetlands and lakes, potentially impairing ecosystem function and impeding biological recovery(Alpers et al. 2005 ; Alpers et al. 2016 ; AQUAMIN Steering Group 1996 ). Over the past three years, our group has made significant headway on assessing the impacts of contaminated tailings to freshwater ecosystems in the field and in the laboratory (Chapman, et al. 2020 ; LeBlanc 2019a ; LeBlanc 2019b ; Gaudet 2022 ). We have shown that sediment and water at these sites are acutely toxic to sensitive aquatic invertebrates, and that more tolerant aquatic invertebrates and amphibians in impacted wetlands bioaccumulate significantly higher Hg and As concentrations than those from reference wetlands. We also examined As and Hg concentrations in adult dragonflies and their juvenile aquatic nymph counterparts, and have shown that the aquatic nymphs accumulate significant As and Hg concentrations underwater (LeBlanc 2019b ). When the nymphs emerge as adults, the adult dragonflies tend to have consistently elevated Hg concentrations and will have elevated As concentrations at sites with high levels of As in sediments. Those invertebrates present an ecosystem risk because top predators, both underwater (fish, diving birds) and terrestrial (bats, birds, insect-eating mammals) will be exposed to significant Hg and As through consuming those. Because Hg and As can be transferred up food webs, economically-valuable species and species at risk likely are impacted which complicate future environmental assessments and raises liability issues for Crown land. Natural recovery strategies (i.e., leaving these sites alone) have clearly not been successful, with many legacy gold-mine-tailing wetlands still exhibiting severe contamination issues more than 100 years after the historical gold rushes ended. Nova Scotia wetland and shallow-water sites impacted by legacy gold-mine tailing material are frequently situated in remote areas and are near high-value freshwater ecosystems supporting economically valuable fisheries and recreation sites. Dredging to remove contaminated material would be prohibitive at this provincial scale, and can cause a re-release of As and Hg back into freshwater ecosystems (DeSisto et al. 2017 ). Large remediation projects involving excavation and off-site treatment would also consume significant amounts of energy and emit large quantities of greenhouse gases (Hou et al. 2023 ). The government of Nova Scotia recently estimated the cost of remediation of the Crown land portion of two legacy gold mining sites at $ 60 million CAD (Intrinsik Corp, et al. 2019a, 2019b). High-level conceptual closure plans prepared for these two sites included recommendations to move contaminated tailings on land to containment cells, while contaminated wetland areas be capped due to the risk associated with disturbing wetland sediments. Capping of freshwater sediments can provide cost-effective isolation and exposure-pathway elimination. Conventional (passive) caps or low permeability liners, normally uses clean, neutral materials like sand, silt, clay, and crushed-rock debris and rely on containment rather than treatment (Zhang et al. 2016 ). These materials are often easy to find at a moderate cost. However, these types of caps require a thick layer of material, often at least 50 cm to be effective in physical isolation of contaminated sediment. In shallow wetland environments commonly associated with Nova Scotia sites, reductions in wetland flood storage capacity and depth of the overlaying water needs to be considered. There is a need for innovative, resource-efficient in-situ remediation strategies that can cost-effectively manage the risk of both As and Hg in wetland sediments while retaining and improving wetland functionality (Intrinsik Corp, et al. 2019a, 2019b). The thickness, function, and longevity of traditional, passive isolation caps can be improved through the use of reactive capping materials, which ideally reduce contaminants’ mobility, toxicity, and bioavailability and thereby offer both containment and treatment of contaminated sediment. Olsta ( 2007 ) states that a 12 mm (0.5 in) thick reactive mat can theoretically replace 1 m (3 ft) of sand or soil. Reactive materials that have been used in cappings of contaminated sediments include activated carbon, biochar, composts, organoclays, calcite, zeolite, bentonite, apatite, biopolymers, and zerovalent iron (ZVI) (Zhang et al. 2016 ). Unfortunately, common reactive soil amendments, especially organic amendments used to reduce bioavailability of some metals can significantly increase the mobility and bioaccessibility of As in wet soils and sediment, leading to elevated human and ecological risks (Cerqueira et al. 2022 ; Meunier et al. 2011; Saunders et al. 2011 ). Consequently, using organic amendments and fertilizer additions, which is common practice for phytostabilization projects elsewhere, may not be a feasible stand-alone option for historical gold-mining tailing sites with elevated As and fluctuating water levels (Meunier et al. 2011). Knox et al., ( 2011 ) studied multiple amendment active caps (MAACs) for the remediation of contaminated sediments, which consist of a mixture of chemically active amendments combined with sand or other neutral materials. They found that phosphate, zeolite, bentonite, and organoclays individually or mixed with another active or neutral materials can stabilize metals and nonpolar pollutants (e.g., PAHs), and addition of a small amount of bentonite (e.g., 10%) to MAACs can improve erosion resistance and metal sequestration capacity. Zeolites are crystalline hydrated aluminosilicates of alkali and alkaline earth elements with a very high cation exchange capacity of up to 6 mmol(eq)/g. Due to the high cation exchange capacity, natural zeolites are capable of demobilising large amounts of cationic pollutants by sorption (Jacobs and Förstner 1999), but zeolite has a very low affinity for anionic compounds because of negative surface charges. Jeon et al., ( 2009 ) tested iron (III) coated zeolite for As(V) removal and found that As(V) was combined with iron oxyhydroxide onto the zeolite by complexation, but the adsorption capacity was insufficient to ameliorate high contamination levels. Li et al., ( 2018 ) suggested that nanoscale zero-valent iron (nZVI) with a high anionic adsorption capacity and unique core-shell structures (Mu et al. 2017 ), combined with zeolite may overcome the deficiencies of zeolite for treatment of As. nZVI is also relatively low-cost, in comparison with other metallic nanoparticles (Arshadi et al. 2017 ; Gil-Díaz et al. 2017 ), and is an effective reducing agent for metal (loid)s (Arshadi et al. 2017 ). ZVI treatment has been shown to lower the toxicity of As and Hg through converting oxidized elements to less mobile and toxic forms. Additionally, all nZVI particles consists of a zero-valent iron (Fe0) core and an iron oxide shell structure. The nature of the iron oxide shell on the Fe0 core of nZVI particles has an important impact on the inorganic contaminant removal performance of nZVI. It has been show that the adsorption of inorganic species by nZVI was predominantly mediated by this oxide shell (Mu et al. 2017 ). Mu et al. ( 2017 ) reported that the oxide shell might consist of a single phase or be made up of several phases (such as wüstite (FeO), magnetite (Fe3O4), maghemite (γ-Fe2O3), hematite (α-Fe2O3), and goethite (FeOOH)). In an earlier study: (Chapman et al. 2020 ), we explored whether nZVI added to two different contaminated wetland sediments could reduce Hg and As mobility and toxicity to two aquatic invertebrates; burrowing mayflies ( Hexagenia spp ) and Chinese mystery snails ( Cipangopaludina chinensis ). Total Hg and As concentrations in overlaying water above both contaminated sediments were reduced by at least 75% and 88% respectively when treated with nZVI. In the first sediment, juvenile snail survival increased from 75% in the untreated sediment to 100% in all nZVI treatments. The 2% nZVI treatment level was the only one with surviving mayflies (33%) and growth of juvenile snails. No snails or mayflies survived in the second sediment with higher Hg and As concentrations regardless of nZVI treatment level. It is possible that this is because nZVI tends to aggregate rapidly in water, which reduces its adsorption capacity significantly. In order to prevent this aggregation of nZVI, different types of clay minerals have been successfully used to support nZVI, such as zeolite (Kong et al. 2016 ; Li et al. 2018 ; Wang et al. 2010 ). Zeolite-supported nZVI has mostly been synthesized in laboratories using liquid phase reduction and ion exchange procedures (Kong et al. 2016 ; Li et al. 2020 ; 2018 ; Wang et al. 2010 ). In short, this involves first blending zeolite with ferric chloride or FeSO 4 7H 2 O in different ratios with water, pH-adjusted and stirred vigorously. Next, aNaBH 4 solution is added to ensure adequate reduction of Fe(III). Finally, the zeolite-nZVI composite is separated from the mixture solution by magnet, centrifuge, or vacuum filtration, washed with ultrapure water to remove soluble impurities, and then usually stored in vacuum before use. This process is time consuming and requires laboratory-grade equipment and chemicals in large quantities, making this impractical for large-scale field application. Commercially-available sources of nZVI and zeolite and simpler production methods for blends of these ingredients would be preferable for bulk distribution in wetland areas. For example, Liu et al., ( 2017 ) tested an iron-oxide coated zeolite for the pollution control of river sediments. The iron-oxide coated zeolite in this study was prepared by a simple mixing of moistened iron oxide powder with deionized water and zeolite. This iron-oxide coated product showed capacities to sorb ammonia, phosphates and sulfides from sediments. It may therefore be possible to coat zeolite with nZVI particles in a similar procedure as outlined for iron-oxides by Liu et al. ( 2017 ), but this has not yet been investigated. NANOFER STAR is a commercially available prepared dry air-stable nZVI powder, where the surface of iron nanoparticles is stabilized by a thin layer of iron oxide, which prevents immediate oxidation in contact with atmospheric oxygen. The nanoparticles of this product are in form of clusters and agglomerates and it is therefore necessary to “activate” the nanoparticles by dispersing them and eroding the iron oxide layer in an aqueous suspension (slurry). According to the manufacturer, the highest reactivity of this slurry was following 48 hours after suspension of powder in water (NANOIRON 2010 ). NANOFER STAR product becomes very reactive in water environment, hydrogen gas is produced during reaction of the product with water and Fe(0) nanoparticles are transformed to iron oxides and hydroxides. Therefore, it is necessary to apply the slurry as soon as possible following the 48-hour activation period. Zeolite has not been added to this product, but if incorporated during the mixing and activation period, we hypothesize that nZVI will coat the zeolite, which in turn could lead to reduced agglomeration and enhanced reactivity for Hg and As contamination. The objectives of this study were: 1) to test a simple method for coating the NANOFER STAR nZVI on zeolite, and 2) to test this nZVI coated zeolite blend as a reactive barrier under a protective capping layer of sand and clays, and as a component of the protective capping in an “active cap” for limiting the mobility, bioaccumulation and toxicity of Hg and As in wetland sediment impacted by legacy gold mine tailings. Methods Contaminated sediment collection and control sediment preparation The contaminated wetland sites for this study were the Muddy Pond (MP) wetland, located within the Waverley historical gold mining district and the Old Stamp Mill (OS) wetland, located within the Montague historical gold mining district, both in the Halifax Regional Municipality, Nova Scotia, Canada. We collected 18 10-cm deep soil cores (diameter 7–8 cm) from each site, which were immediately inserted into heavy-duty zip-lock bags. Care was taken to ensure minimal disturbance of the sediment and contact with air. The sediment was stored cold (7˚C), in vacuum, and in the dark until the start of the test. An artificial control sediment was prepared with the goal to mimic the texture and pH of the tailing sediments. This consisted of 90% fine sand (< 180 um), 5% peat (sieved to < 1mm) and 5% clay. CaCO 3 was added to the sediment in order to increase pH to 6–7. Particle size distribution of the wetland and control sediments was determined using laser diffraction particle size analysis by Loring Tarcore Labs Ltd for clay content (Table 1 ). Organic matter was determined by percent loss on ignition (% LOI). Dry sediment material was weighed and placed in a muffle furnace at 550°C for exactly 4 h. Samples were then re-weighed and % loss on ignition was calculated from sediment weight differences between 60˚C and 550°C (Table 1 ). Table 1 Sediment type and properties (metal(oid)s, sulphur, organic matter content as determined by loss on ignition (LOI), clay content, and pH over overlaying water at test start) Sediment type/properties Muddy Pond (MP) Old Stamp Mill (OS) Control/artificial sediment (C) CCME PEL (mg/kg) As (mg/kg) (n = 18) 94 211 ± 25 437 788 ± 560 0.3 ± 0.2 17 Hg (mg/kg) (n = 18) 41 ± 13 91 ± 16 0.012 ± 0.004 0.486 Pb (mg/kg) (n = 18) 716 ± 224 710 ± 112 5.0 ± 0.3 91.3 Cu (mg/kg) (n = 18) 54 ± 15 478 ± 71 0.7 ± 0.1 197 Ni (mg/kg) (n = 18) 133 ± 38 18 ± 2 2.0 ± 0.2 75 Fe (mg/kg) (n = 18) 95 161 ± 24 823 19 728 ± 1513 489 ± 47 43 766 Ag (mg/kg) (n = 18) 1.4 ± 0.6 2.0 ± 0.2 0.008 ± 0.001 1 Zn (mg/kg) (n = 18) 657 ± 140 314 ± 34 1.3 ± 0.1 315 S% (n = 18) 4.7 ± 1.1 0.03 ± 0.02 0.02 ± 0 - LOI% (n = 18) 9.3 ± 2.9 3.8 ± 1.1 6.4 ± 0.6 - % Clay (n = 1) 4.4 4.5 5.0 - pH (overlaying water) (n = 6) 6.9 ± 0.1 6.4 ± 0.1 7.5 ± 0.2 - nZVI/zeolite reactive blend and protective capping material preparation A NANOFERSTAR nZVI (NANOIRON, 2010 ) stock slurry was prepared by blending 50 g of nZVI (with ~ 80% active Fe 0 ) and 200 mL of dechlorinated water for 10 minutes in 1000W blender (NANOIRON 2010 ). Next, 80g of Zeolite (ultrafine < 20µm, 90–92% Clinoptilolite from Heiltropfen®) was added to the nZVI stock slurry and blended for another 10 minutes aiming for a ratio of 2:1 zeolite/nZVI as per Arancibia-Miranda et al., ( 2016 ) and Wang et al., ( 2010 ). The Zeolite/nZVI slurry was then activated by storing for 48 hrs in a sealed plastic amber PET bottle at room temperature. It was observed that following the activation period, the hydrogen production (visually assessed by bulging of the PET bottle) was much greater in the zeolite/nZVI slurry compared with a control slurry containing nZVI only, which indicates higher reactivity in the zeolite/nZVI slurry. The protective capping layer was prepared by mixing fine silica sand (80%), bentonite - montmorillonite (10%) and ultrafine micronized (< 20um) zeolite -clinoptilolite (10%). Samples of NANOFERSTAR nZVI activated slurry, zeolite powder, and nZVI-coated zeolite activated slurry were cold mounted in a polished epoxy resin. The surface morphology, particle size and elemental distributions were then determined using the TESCAN Mira3 LMU scanning electronic microscope (SEM) (maximum resolution up to 1.2 nm at 30 kV combined with an energy dispersive X-ray (EDX) analysis) in the Saint Mary’s University Electron Microscopy Centre. Experimental design The multi-phase experimental design (Fig. 1 ) included a total of 54 1-L beakers, with 18 beakers containing 100-mL of each sediment type (MP, OS and control). Within those 18 beakers, 6 replicates were left untreated while the other 12 received two different treatments: “A” or “B” (each with 6 replicates) as outlined below. Treatment “A” consisted of 5% nZVI coated zeolite slurry added to the top of each sediment based on wet weight of sediment, then 50 ml (1cm) protective capping material was added on top of this. The 5% nZVI coated zeolite concentration equates to 30 g nZVI coated zeolite per kg of sediment. This concentration was selected as Li et al. ( 2018 ) found that arsenic, cadmium and lead in soil samples were immobilized at this concentration with their zeolite-supported nZVI. Treatment “B” consisted of a 100 ml (2-cm) layer of protective capping material mixed with the nZVI coated zeolite at 5% of wet weight of capping added to the top of each sediment. All beakers then received 650 ml (9-cm water column above the top of the sediment) of specially prepared “ Hyalella azteca culture water”. This culture water was prepared based on federal government ecotoxicology standard protocols (Environment and Climate Change Canada 2017 ). Beakers with sediment and water were then allowed to settle for 10 days before starting the Hg bioaccumulation experiment. Hg bioaccumulation experiment Large-bodied invertebrates used for bioaccumulation studies (Fig. 1 ) included Amano shrimp ( Caridina multidentate ) and spike-topped apple snails ( Pomacea bridgesi ). These were selected for their relative tolerance of contamination and large individual biomass, allowing tissue analysis of Hg. Three replicates of each treatment #1–3 received one snail each, and three replicates of each treatment #4–6 received one shrimp. Snails and shrimp were left unfed for 5 days; the duration of the test, which was determined by a preliminary experiment showing that shrimp could not survive past 7 days in jars with clean silica sand without feeding. Following the 5-day exposure period, the shrimp and snails were retrieved and allowed to depurate in clean test water for 24 hours. Snails and shrimp were euthanized by rapidly freezing them at -20C˚. All invertebrates were then individually weighed, and dried at ≤ 60℃, then weighed again to determine moisture content. After removing snails from their shells, dried whole-body tissues were ground to a powder using an Retsch MM 400 ball mill mixer prior to analysis. Invertebrate ecotoxicity experiments Five days after the snails and shrimp were removed from the beakers, a 14-day toxicity test with Hyalella azteca was initiated (Fig. 1 ) in the same beakers, following Environment and Climate Change Canada ( 2017 ) ecotoxicology protocols. Briefly, ten juvenile Hyalella (age 1–7 days) were placed in each of 5 replicates/treatment for a 14-day exposure. Lights were set to 500–1000 lux on a 16-h light, 8-hr dark cycle, and juvenile Hyalella were fed 3x/week. At the beginning of the test, the overlaying water in the beakers containing control sediment had elevated ammonia levels above recommended levels for Hyalella, so 180 mL of water in each jar was replaced with fresh test water 3x/week. At the end of the test, the final dry weight of Hyalella azteca was noted as well as % survival in each replicate. Immediately following the toxicity test with Hyalella azteca , a toxicity test with Daphnia magna was initiated (Fig. 1 ) following ecotoxicology protocols Environment and Climate Change Canada (2014). Water (150 ml) from each 1L beaker used in the Hyalella azteca toxicity test was decanted into a 250-mL beaker and 10 Daphnia magna neonates (ages < 24h) were added for a 48-hr exposure test. At the end of the exposure, the number of mobile/surviving daphnids were counted. Analysis of Hg, As, Fe and other elements Before adding treatments to sediments and test water, samples of contaminated and artificial control sediment were collected from each beaker as well as the protective capping material. Sediments and protective capping material were dried at a low temperature (40–60℃), pulverized, and then shipped to Bureau Veritas Minerals Laboratory (Vancouver) for modified aqua regia digestion (1:1:1 HNO3:HCl:H2O) and analysed using ICP-MS for ultra-low determination. Quality controls included sample blanks, duplicates, reagent blanks and Bureau Veritas Certified Reference Materials (OREAS262 and DS11). Temperature, pH, conductivity, dissolved oxygen, ammonia and alkalinity levels were monitored regularly in the overlaying water of each replicate to ensure acceptable levels for test organisms. At the end of the Hg bioaccumulation test (Day 15), and then again at the end of the Hyalella azteca ecotoxicity tests (Day 34), two 100-ml water samples were obtained from each beaker for total (unfiltered) and dissolved metal(oid) analysis (filtered through 0.45 um Teflon DigiFILTER), and all samples preserved with 1% v/v ultrapure nitric acid. The 100-mL filtered and unfiltered samples were analysed for metal(loids) on ICP-MS by the Analytical Services Unit Laboratory at Queens University. Quality controls included CRMs (EU-H) and in-house reference water samples, all within the 10% range of uncertainty, stock standards, 10% duplicate samples and blanks. Additionally, at the end of the Hyalella azteca toxicity tests (day 34), 2 mL porewater was non-destructively sampled from the sediments using syringe-driven Micro-Rhizon© porous samplers. Porewater was collected from three replicates of each treatment, for a total of 27 samples and preserved with 1% v/v ultrapure nitric acid. As the porous part of MicroRhizons samplers have a mean pore size 0.15 µm, additional filtering of the samples before analyzing was unnecessary. Given the small porewater sample volumes, we were only able to analyse Hg via a Direct Mercury Analyzer (DMA) at Saint Mary’s University. Results for dissolved Hg analysis were within the 10% range of uncertainty, and each analytical run also included blanks, standard calibration check samples and 10% duplicates. Filtered and unfiltered water samples, snail and shrimp tissues were analysed for Hg on the Milestone Direct Mercury Analyzer DMA 80.3 at Saint Mary’s University. Quality control for each analysis run included blanks, calibration solutions, 10% duplicates and reference materials (National Resources Canada NRCan TORT-3 and DORM-4 for tissue analyses, in-house spiked reference water samples for water). Data analysis Data were assessed for normality and equal variance. Since our datasets were non-parametric and varied between 3–6 replicates, the different treatments were assessed using Kruskal-Wallis (K-W) One Way Analysis of Variance on Ranks with pairwise multiple comparison procedure. Results nZVI and nZVI coated zeolite characteristics SEM scans of the nanoscale zerovalent iron particles in activated NANOFER STAR slurry indicated an approximate grain size of 50–100 nm, and that particles were highly agglomerated (Fig. 2 b). In contrast, although some agglomeration was still observed, when blended with zeolite during the activation phase, nZVI particles in the NANOFER STAR slurry appeared more evenly spread across the zeolite surface, and appeared with less agglomeration than for nZVI alone (Fig. 2 c). The surface elemental distributions of the three different materials are also illustrated in Fig. 2 . The atomic weight composition for the zeolite- nZVI blend (2c) suggests elemental distributions similar to both zeolite and nZVI (Fig. 2 a and b). However, nZVI particles appear more concentrated in cracks or pores on the zeolite particles surface as illustrated in the different spectras for the zeolite-nZVI blended material in Fig. 2 c. Sediment and porewater characterization Of the three different sediments, MP had the highest organic matter content (% LOI) at 9.3% followed by the artificial control sediment at 6.4% and then OS at 3.8%. Clay content was similar between the different sediments ranging between 4.4-5% (Table 1 ). The lowest pH was measured in the overlaying water above the OS sediment at 6.4, while the overlaying water pH for artificial control sediment was 7.5. Sulphur content was higher in MP sediment at 4.7% compared with OS (0.03%) and control sediment (0.02%). Total As and Hg concentrations in both sets of contaminated sediments (Table 1 ) exceeded Canadian Council of Ministers for the Environment (CCME) sediment Predicted Effect Level (PEL) guidelines for As (17 mg As/kg) and Hg (0.486 mg Hg/kg). Total sediment As concentration in the untreated MP sediment was very high with an average of 94,211 ± 25,437 mg As/kg sediment (Table 1 ), while the OS sediments were lower, but still elevated (788 ± 560 mg As/kg). This trend is reversed for Hg, with OS sediments (91 ± 16 mg Hg/kg) having higher concentrations than the MP sediments (41 ± 13 mg/kg). Other metals exceeding CCME PEL in the sediments included Pb, Ni, Fe, Ag and Zn in MP sediment and Pb, Cu, Ag and Zn in OS sediment (Table 1 ). The prepared artificial control sediments had metal(oid) concentrations below the CCME PEL guidelines (Table 1 ). Porewater Hg concentrations (Fig. 3 ) in Old Stamp Mill (OS) sediments treated with “B”; OSB (0.2 ± 0.1µg/L) was significantly lower than for untreated OS (0.9 ± 0.2 µg/L) (K-W p = 0.004). The porewater Hg concentration from OSB was similar to porewater Hg concentrations in the artificial control sediment. There was no significant difference in Hg porewater concentrations between treated and untreated Muddy Pond (MP) sediment, or between any of the artificial control treatments. Total and Dissolved As, Hg and other metal(loids) in overlying water Total and dissolved As, Hg and other metal(loids) in the overlaying water above treated and untreated sediments was measured at day 15 (end of the Hg bioaccumulation test, Fig. 1 ) and then again at day 34 (end of the Hyalella azteca sediment ecotoxicity test, Fig. 1 ). In general, total concentrations of As and Hg in the overlaying water decreased (Figs. 4 and 5 ) from day 15 to day 34 (Fig. 1 ) as can be expected when particle bound As and Hg settle out in the beakers. On day 34, between 75–100% of total As in the overlaying water was in the dissolved fraction for all sediments and treatments, with one exception for overlaying water in MPB with 54% of As in dissolved form. Of the total Hg concentrations, between 50–100% in the overlaying water of all sediments and treatments on day 34 was in dissolved form. Average total [As] in the overlaying water above untreated MP sediment was very high on Day 15 (Fig. 4 ) at 1482 ± 605 µg As/L and remained relatively high on Day 34 at 1088 ± 197 µg As/L. The Day-34 total [As] was significantly higher (K-W, p = < 0.001) than what was found in the overlaying water above controls (1 ± 0.2 µg As/L), and with the B treatment (3 ± 0.8µg/L), but not significantly different than what was found with treatment A (17 ± 7µg/L). It is noteworthy that the treatment B drastically decreased total [As] in the overlaying water above MP sediment to below the CCME guideline value for protection of aquatic life at 5 µg/L. Average total [As] in the overlaying water above the untreated OS sediment on day 15 was 199 ± 267 µg/L (Fig. 4 ). This decreased to 71 ± 36 µg/L on day 34. Even so, this was significantly higher (K-W, p = < 0.001) than the total concentrations in the overlaying water above control sediment and with the B treatment (2.9 ± 2.7µg/L). Treatment B was able to decrease [As] in overlaying water above OSB sediment to concentrations below CCME guideline values for protection of aquatic life. Total [As] concentrations in OS sediment with treatment “A” measured 9 ± 7µg/L. This concentration was not significantly different (K-W, p = < 0.001) than in the overlaying water of untreated sediment. The highest total [Hg] was found in the overlaying water above the untreated OS sediment on Day 15 (2.4 µg/L ± 1.4). This concentration was significantly higher (K-W, p = 0.016) than for those in controls, as well as treatment A (0.87 µg total Hg /L ± 0.96) and B (0.28 µg Hg /L ± 0.13) (Fig. 5 ). On Day 34, the total [Hg] in the overlaying water above the untreated OS sediment had decreased to 0.42 µg Hg/L ± 0.22, indicating the settling out of particulate Hg. This total Hg concentration was no longer significantly different than concentrations found in the overlaying water above controls and treatment A (0.22 µg/L ± 0.04) and B (0.26 µg/L ± 0.09). The overlaying water for the untreated Muddy Pond (MP) sediment measured 0.73 µg total Hg /L ± 0.86 on Day 15 and was not significantly higher than overlaying water Hg concentrations in the controls and treatment A and B (Fig. 5 ). On Day 34, the overlaying water Hg concentrations from the untreated MP sediments decreased to 0.2 µg/L ± 0.0. It should be noted that total and dissolved Hg in the overlaying water above treated and untreated controls on Day 15 and Day 34 ranged between 0.2–0.4 µg/L, and this is higher than the CCME guideline value for protection of aquatic life at 0.026µg total Hg/L (Fig. 5 ). On day 15, 100% of the total Hg in the overlaying water in both treated and untreated controls was in the dissolved fraction. The fraction of dissolved Hg remained high (100%) in the untreated control on Day 34, but decreased in treated control sediment to 61% (treatment A), and 52% (treatment B). Total overlaying water concentrations of other metal(loid)s which exceeded CCME PEL’s included Pb, Cu, and Fe in some cases. In summary, total overlaying water Pb concentrations on Day 15 in untreated control (6 ± 6 µg/L), MP (8 ± 11 µg/L) and OS (32 ± 40µg/L) sediment all exceeded the CCME CWQG Guideline of 1 µg/L. However, on Day 34, total overlaying water Pb concentrations in untreated control (0.5 ± 0 g/L) and MP (0.5 ± 0 g/L) sediments dropped to below the CCME guidelines, while untreated OS overlaying water concentrations (3 ± 2µg/L) still remained above CCME guidelines. On Day 15, the overlaying water total [Cu] in untreated controls was 2 µg ± 0.3, while untreated MP was 2 µg ± 0.04, MPA 10 µg ± 9 and MPB 16 µg ± 13. Untreated OS sediment measured 72 µg ± 36, OSA 13 µg ± 10 and OSB 9 µg ± 5. On Day 34, the overlaying water [Cu] for untreated controls, untreated MP and MPA all reported levels of Cu below the CCME CWQG guideline value (2 µg/L). MPB reported a slightly higher concentration at 4 µg ± 2. Overlaying water total [Cu] for untreated OS on day 34 was 28 ± 18µg/L, exceeding the CCME guideline value, and was significantly reduced for OSA (by 85%) and OSB (by 81%) treatments. Overlaying water total [Fe] on Day 34 in untreated MP, MPA and MPB was higher than those on Day 15. Only the MPB (614 ± 607 µg/L) overlaying water samples on Day 34 exceeded CCME PEL guideline of 300µg/L, out of which 8% was in the dissolved form. Bioaccumulation of Hg Amano shrimp exposed to OS sediments (Fig. 6 ) had lower Hg concentrations in the beakers with treatments A (0.7 ± 0.3 mg Hg/kg) and B (0.08 ± 0.03 mg Hg/kg) compared to those from untreated OS sediments (6.0 ± 5.0 mg Hg/kg), and similar to those found in the control sediments (0.07–0.08 mg Hg/kg). Amano shrimp exposed to MP sediments (Fig. 6 ) had moderately elevated Hg in untreated MP sediments (2.9 ± 3.5 mg Hg/kg) and appeared to have lower Hg concentrations in the beakers with treatment A (0.8 ± 0.8 mg Hg/kg) and B (0.2 ± 0.3 mg Hg/kg), but those were not statistically significant. The apple snails ( Pomacea bridgesi ) did not have statistically significant differences in Hg tissue concentrations between any of the contaminated sediment, treated sediment or control sediment, and we observed the snails in the beakers often had their operculum closed during the exposure. Ecotoxicity Sediment Test with Hyalella azteca Hyalella azteca 14-day survival in untreated control sediment was satisfactory with an average survival of 94 ± 9% (Fig. 7 ), which is in accord with the ECCC Hyalella test method requiring > 80% survivability in control treatments (Environment and Climate Change Canada 2017 ). H. azteca also had high survivability in control sediment with treatment A (96 ± 5%) and treatment B (88 ± 4%). The average individual dry weight (0.4 ± 0.1mg) of control H. azteca did not change significantly between treated and untreated controls. H.azteca 14-d survivability in untreated OS sediment (56 ± 22%) was lower than for the control sediments. However, 14-day survivability significantly increased in treatment A (94 ± 6%, p = 0.006) comparable to survival in control sediments. Survival in treatment B also appeared higher at 90 ± 7%. However, this apparent increase in survival was not statistically significant. The average individual dry weight of H. azteca (0.1 ± 0.03 mg) in OS sediments was not significantly different between treated and untreated OS, but was significantly lower than those for control H. azteca (p = 0.008). The 14-d survivability of H. azteca in untreated MP sediments was very low (14 ± 9%), with only 7 individual amphipods surviving to the end, as opposed to 45–50 individuals in all other treatments. There were significant improvements in 14-d survivability for MPA (96 ± 5%, p < 0.001) and MPB (90 ± 10%, p < 0.001) treatments. There was no significant difference between the dry weights of H. azteca exposed to treated and intreated MP sediments, or between MP sediments and controls, but the low survivability of the untreated MP amphipods made it difficult to assess significance. Ecotoxicity Water Test with Daphna magna The 48-h survival of Daphnia magna in overlaying water taken from above untreated sediments and treated control (98%±5) were similar and is acceptable in accordance with the ECCC standard requiring > 90% survivability in control treatments (Environment Canada 2014). The 48-h survival (62%±34) of D. magna in water from untreated OS jars varied considerably among replicates (Fig. 8 ). The 48-h survivability improved for OSA (94%±9) and OSB (96%±6) treatments although the improvement was not statistically significant (p = 0.06). The 48-h D. magna survival for water from the untreated MP sediment (84%±9) was lower than those for the MPA (98%±5) and MPB (96%±5) sediment treatments, with MPA being significantly higher (p = 0.02). Discussion The customized reactive amendment component that was placed on two different contaminated sediments, either below (treatment A), or within (treatment B) the thin protective capping matrix in this study is a novel simple blend of commercially available NANOFER STAR nZVI and fine-grained zeolite (clinoptilolite). When the activated NANOFERSTAR slurry was mechanically blended with zeolite, nZVI particles still agglomerated due to Vander Waals forces and magnetic interactions (He et al., 2007), but the agglomeration of nZVI particles was reduced and they appeared more evenly distributed on the surface of zeolite. This is similar to what was found by Li et al., ( 2018 ) who used more labour and resource intensive liquid phase reduction and ion exchange procedures to synthesise their zeolite supported nZVI product. Reduction in agglomeration of nZVI and dispersion of nZVI particles of zeolite in their study was more pronounced than in ours. They noted that nZVI particles were homogeneously dispersed on the surface of zeolite, with no obvious agglomeration observed. This more even distribution of nZVI particles on the zeolite is likely due to the liquid phase reduction and ion exchange procedures used to synthesize the zeolite supported nZVI. In our study mechanical mixing and introduction of zeolite during the activation stage of the nZVI was not as effective in dispersing the nZVI on the zeolite. Nevertheless, hydrogen production was much greater in the zeolite/nZVI slurry compared with a control slurry containing nZVI only, indicating higher reactivity in the zeolite/nZVI slurry blend than in the nZVI slurry alone. SEM images also confirm that coating of NANOFERSTAR nZVI on zeolite (Fig. 2 C) using simple mechanical blending during the activation phase prevented nZVI from agglomerating to some degree, which increased the specific surface area and reactivity of nZVI. Can Treatment A and B reduce migration of As, Hg, and other metals from contaminated sediments? Both treatment A and B reduced total [As] in the overlaying water above MP and OS sediments but the Treatment B (zeolite-nZVI blend mixed in with a 2cm layer of protective capping) was more successful. Treatment B beakers exhibited decreased total [As] in the overlaying water above MP sediments by 99%, and by 96% in OS sediments, bringing total [As] to below the CCME guideline value for protection of aquatic life (5 µg/L). No other studies to the best of our knowledge have attempted to treat As and Hg contaminated wetland sediment with a zeolite-nZVI customized thin reactive capping blend, or zeolite supported nZVI. However, studies have investigated separate components of our blend for direct amendment of contaminated soil and sediment. For example, Li et al. ( 2018 ) evaluated the bioavailability of metals in aquatic sediments amended with zeolite and found that As porewater concentrations were reduced by 71% after 24 hours. This is in contrast to findings by Kang et al. ( 2016 ) who discovered higher As concentrations above sediments amended with zeolite only. Since zeolite has a very low affinity for anionic compounds due to negative surface charges (Jacobs and Förstner, 1999), this is not surprising. Nanoscale zero-valent iron (nZVI) on the other hand, has a high anionic adsorption capacity (Mu et al., 2017 ). Gil-Díaz et al., ( 2017 ) found that an nZVI dose of only 5% to soil lead to a 70% decrease of exchangeable As. This is similar to a previous study completed by our laboratory group (Chapman et al. 2020 ), where an 8% dose of nZVI was able to reduce As concentrations in the overlaying water above As-contaminated sediment by 88%. In our current study, in Treatment B where zeolite was combined with nZVI, not only did the total [As] decrease more drastically, but the fraction of dissolved As (potentially more toxic forms) also decreased from 100–54% above MP sediments. It appears that our zeolite-nZVI blend and a protective capping layer is more successful in reducing As migration from sediments than nZVI or zeolite alone. This is supported by several batch-testing studies assessing zeolite and iron products in As-spiked water. Jeon et al. ( 2009 ) investigated the sorption characteristics of arsenic (As(V)) on iron-coated zeolite (ICZ) through batch studies, and they found that As(V) was completely removed within 30 min in a concentration of 2 mg/l, with a 100 g/l dose of ICZ. The adsorption capacity of ICZ for As(V) was 0.68 mg/g. However, arsenite (As(III)) is more prevalent in anoxic or acidic waters common to wetland areas than As(V). As(III) is also more mobile and toxic than arsenate (Sealy, 2011 ). Li et al. ( 2018 ) used batch tests to investigate zeolite-supported nanoscale zero-valent iron for adsorption of As(III) in aqueous solution. They found that the maximum adsorption capacity for their zeolite supported nZVI was 11.52 mg As(III)/g, which was much higher than that of zeolite alone. It was suggested that As reacts with zeolite supported nZVI through adsorption mechanisms including electrostatic adsorption, ionic exchange, oxidation, reduction, co-precipitation, and complexation. Of the two treatments in this study, Treatment B also appeared most effective in preventing Hg release into the overlaying water and porewater from sediments. On day 15, total [Hg] in the overlaying water above Treatment B MP sediment (MPB) had been significantly reduced by 65%, and above “B” treated OS sediment by 88%. On day 34, [Hg] in the overlaying water in untreated MP and OS beakers were still elevated compared with CCME guideline values but did not differ significantly from total [Hg] found in controls or treatments. This was also the case for dissolved Hg porewater concentrations in treated and untreated MP sediments. However, both Treatments A and B significantly reduced dissolved [Hg] in OS sediment porewater. The dissolved porewater [Hg] was reduced by 35% in OSA and by 73% in OSB. This is similar to findings by Lewis et al. ( 2016 ) who investigated nZVI Hg remediation in wetland sediment in a mesocosm set up. They analysed porewater for total Hg and methyl Hg and found that total concentrations decreased in porewater treated with nZVI by approximately 27% and methyl mercury by 42%. They hypothesized that ZVI likely decrease the MeHg concentration via adsorption and not demethylation or inhibition of Hg(II) methylation. However, Gil-Díaz et al. ( 2017 ) found that a dose of 5% of nZVI to contaminated soil did not significantly reduce exchangeable Hg. A higher dose of nZVI (10%) was necessary in their study to achieve reductions of exchangeable-Hg, between 63 and 90% depending on the type of nZVI and soil. Our 73% reduction of dissolved Hg porewater concentrations on addition of treatment “B” with lower nZVI concentrations than in the Gil-Díaz et al. ( 2017 ) study could be due to zeolite’s additional adsorption of Hg. Zeolites are proven ion exchange materials where the indigenous (typically sodium) charge balancing cations are readily exchanged with metal cations in solution (Wang et al. 2010 ). Both Treatments A and B significantly reduced overlaying water Pb and Cu concentrations above contaminated sediments by at least 77% (A) and 76% (B) for Pb and, 90% (A) and 80% for Cu, to below the CCME guideline values. This is slightly less than for Zhang Xin et al. ( 2010 ) who reported 98.8% reduction of Pb(II) in electroplating waste water by using synthesized Kaolin-nZVI.. In OS sediment beakers, there was no significant difference in total [Fe] between untreated and treated OS sediments, but Treatment B had significantly higher total [Fe] in overlaying water above MPB sediments after 34 days (614 ± 607 µg/L), exceeding CCME guideline value (300 ug/L). nZVI can oxidize to Fe 2+ and Fe 3+ rapidly in contact with water and sediment, and very high concentrations of these ions are produced over a short time. This is concerning as even though organisms can tolerate high Fe concentrations, NANOFERSTAR nZVI has been found toxic to Daphnia magna at concentrations exceeding 0.5 mg/L (Keller et al., 2012). However, the literature is mixed on this, because Yoon et al. (2018) found no acute response of Daphnia magna when exposed to nZVI with water Fe concentrations > 100 mg/L. It is noteworthy that our toxicity tests with treated control sediment confirmed that although concentrations of Fe in the overlaying water increased for both Treatments A and B, we did not observe any toxic responses in invertebrates (14 days for Hyalella azteca and 48 hours for Daphnia magna ). The higher total [Fe] in overlaying water above Treatment B in MP sediments (MPB) also was not enough to produce a toxic response in these organisms. This could be because only 8% of the total water [Fe] was dissolved, indicating a lower potential toxicity. It is highly recommended that future assessments of ZVI in-situ treatments for wetland sediments include dissolved as well as total Fe concentrations in overlaying water, as this may clarify some of the confounding results in the literature. In general, Treatment B appeared more effective in reducing mobility of metal(oids) from the contaminated sediments to overlaying water compared with Treatment A. It is hypothesized that this could be due to more effective isolation with the 2-cm thick Treatment B capping compared with the 1-cm thick Treatment A capping. In addition, the active ingredients (zeolite/nZVI) had more contact with the overlaying water in Treatment B because the active ingredients were mixed in with the protective capping as opposed to layered underneath as in Treatment A, and which may have contributed to more effective adsorption of Hg and As in the overlaying water. Can Treatment A and B reduce bioaccumulation of Hg in invertebrates? Total concentrations of Hg in Amano shrimp exposed to OSB sediment was significantly reduced by 99% from 6 mg Hg/kg in untreated OS sediment to 0.08 mg Hg/kg, which were so low that those were comparable with Hg concentrations in the control shrimp treatments (0.07–0.08 mg Hg/kg).In the field, LeBlanc et al. ( 2019 ) reported that invertebrates (dragonfly larvae, damselfly larvae, and aquatic spiders) collected from the Old Stamp Mill contaminated wetland site (OS) has average dry-weight Hg concentrations of 2 mg/kg, while invertebrates from a nearby uncontaminated reference wetland site had 0.17–0.24 mg Hg/kg. It appears Treatment B has reduced the bioavailability of sediment Hg to the point that invertebrate Hg concentrations are well below field values even for invertebrates from reference sites. The much lower Hg concentrations in the OSB shrimp is likely due to Hg binding with zeolite-nZVI leading to reduced bioavailability, but also isolation and reduced exposure to the contaminated sediment provided by the thin protective capping. Treatment A for both OSA and MPA as well as MPB also had much lower Hg concentrations in shrimp, but unfortunately this was not statistically significant. Even so, shrimp from OSA, MPA and MPB beakers had similarly low shrimp Hg concentrations (0.2–0.8 mg Hg/kg) as for OSB shrimp and field data from reference wetlands sites (LeBlanc 2019). Apple snails exposed to treated and untreated wetland sediments did not show a significant difference for Hg concentrations. This is likely due to the fact that this particular snail species ( Pomacea bridgesi ) was able to avoid exposure to sediments by either climbing along the side of the beakers or closing operculum (trap door) over the shell openings. Due to these avoidance techniques, we conclude that this species of snail is not ideal for assessing bioaccumulation of contaminants from sediments. Lewis et al. ( 2016 ) used a different species of snail ( Lymnaea stagnalis ) (which do not have operculum) in laboratory microcosms and was able to confirm that these snails accumulated less MeHg in sediment treated with ZVI. Can Treatment A and B reduce the toxicity of contaminated sediments? The untreated MP and OS sediments were toxic to Hyalella azteca , with only 14% and 56% surviving the 14-day exposure duration respectively. With As concentrations in MP sediment measuring as high as 94,211 mg/kg and in OS sediment; 788 mg/kg, this was expected and similar to what has been found previously. For example, the 10-day LC50 for Hyalella azteca in As spiked sediment has been previously estimated at 532 mg As/kg (Liber et al. 2011 ). A slightly lower 14-day LC50 was reported by Goulet and Thompson, (2018) at 134 mg As/kg. Interestingly, despite the very high As and Hg sediment concentrations in our study, both Treatments A and B were associated with significantly improved survival rates of Hyalella azteca , close to the survival rates observed in controls. Treatment A was slightly more effective in reducing toxicity of sediments to Hyalella azteca than Treatment B for both contaminated sediment types. It is hypothesized that since Treatment A consisted of a reactive layer of concentrated zeolite/nZVI in direct contact with the contaminated sediment, this may have contributed to more effective binding of labile sediment contaminants to the reactive material reducing As and Hg toxicity. Treatment B contained the same zeolite/nZVI mixture, but it was blended in with the protective capping layer so would not have had the same concentration in contact with contaminated sediments. Overlying water concentrations of As in untreated beakers was highest in the overlaying water for untreated MP sediments at 1088 µg/L, and water concentrations of Hg was highest in the water overlaying untreated OS sediments at 0.42 µg/L, which corresponds quite well with toxicity findings of Okamoto et al., (2015), who measured an EC50 of Daphnia magna after 48 hour exposure to be 2400 µg As/L and 0.65 µg Hg/L. Treatments A and B” were associated with increased survival rates of Daphnia magna , with MPA having the highest significance. Survival of Daphnia magna after 48-hour exposure to the overlaying water of untreated MP and OS sediment on day 34 was 84% and 62% respectively. To summarize, our proof-of-concept bench top testing with this reactive zeolite/nZVI amendment blend within and below a thin protective capping shows high potential as an in-situ risk management option for supporting natural recovery of freshwater wetlands impacted by historical gold-mine tailings. The reactive amendments and capping materials are relatively easily processed with commercially-available products. However, to bridge the gap between the laboratory and the real world, it will be necessary to fine-tune component ratios, determine As and Hg adsorption rates of amendment in different sediments, and complete testing using more environmentally realistic mesocosms. It is likely that the nZVI coated zeolite will reduce the thickness of capping needed and improve the overall effectiveness of caps in sequestering contaminants in the sediments, stopping those potentially toxic metal(loids) from migrating into the overlaying aquatic environment. However, concentrations and placement of nZVI in capping materials will have to be carefully assessed before it can be used in the field as the potential environmental impact of nZVI itself is not well known. Declarations Acknowledgements The authors wish to thank Chrisine Moore with Intrinsik Corp for support with this study, as well as Shane Dalton and Anna Murphy for their assistance with experiments/analysis. Funding details: This project was supported by a grant from the Nova Scotia Mineral Resources Development Fund (MRDF) IN-58 (2018). The interns who assisted with the project were supported by SMUworks Work-Study funding from Saint Mary’s University and a Clean Foundation – Environment and Climate Change Canada Professional Internship Program. The Scanning Electron Microscopy (SEM) analyses completed for this study were funded by Atlantic Mining Nova Scotia to Dr. Linda Campbell. Competing interests The authors have no relevant financial or non-financial interests to disclose. Author contributions: Both authors contributed to the study conception, and experiment was designed by E. Emily V. Chapman. Material preparation, data collection and analysis were performed by E. Emily V. Chapman. The first draft of the manuscript was written by E. Emily V. Chapman and Linda M. Cambell commented on previous versions of the manuscript. Both authors read and approved the final manuscript. Ethics approval Ethics approval was not required for this study as it did not involve human subjects or vertebrate animals. In Canada, the Canadian Council on Animal Care do not currently require reviews of invertebrate experiments (with the exception of squid and octopus). We followed best national practices for humane invertebrate care and experimental use, including those of Environment Canada and Climate Change federal experimental protocols. Consent to Participate Not applicable as research did not involve human subjects Consent to Publish Not applicable as research did not involve human subjects References Alpers, Charles, Michael Hunerlach, Jason May, and Roger Hothem. 2005. “Mercury Contamination from Historical Gold Mining in California.” 2005. https://pubs.usgs.gov/fs/2005/3014/. Alpers, Charles N., Julie L. Yee, Joshua T. Ackerman, James L. Orlando, Darrel G. Slotton, and Mark C. Marvin-DiPasquale. 2016. “Prediction of Fish and Sediment Mercury in Streams Using Landscape Variables and Historical Mining.” Science of the Total Environment 571 (November): 364–79. https://doi.org/10.1016/j.scitotenv.2016.05.088. AQUAMIN Steering Group. 1996. “Assessment of the Aquatic Effects of Mining in Canada: AQUAMIN Supporting Document II: Regional Syntheses.” Environment Canada. Https://www.ec.gc.ca/esee-eem/default.asp?lang=En&n=8AF5467D-1. Arancibia-Miranda, Nicolás, Samuel E. Baltazar, Alejandra García, Daniela Muñoz-Lira, Pamela Sepúlveda, María A. Rubio, and Dora Altbir. 2016. “Nanoscale Zero Valent Supported by Zeolite and Montmorillonite: Template Effect of the Removal of Lead Ion from an Aqueous Solution.” Journal of Hazardous Materials 301 (January): 371–80. https://doi.org/10.1016/j.jhazmat.2015.09.007. Arshadi, M., M. K. Abdolmaleki, F. Mousavinia, S. Foroughifard, and A. Karimzadeh. 2017. “Nano Modification of NZVI with an Aquatic Plant Azolla Filiculoides to Remove Pb(II) and Hg(II) from Water: Aging Time and Mechanism Study.” Journal of Colloid and Interface Science 486 (January): 296–308. https://doi.org/10.1016/j.jcis.2016.10.002. Cerqueira, Beatriz, Emma F. Covelo, Sandra Rúa-Díaz, Purificación Marcet, Rubén Forján, José Luis R. Gallego, Lukas Trakal, and Luke Beesley. 2022. “Contrasting Mobility of Arsenic and Copper in a Mining Soil: A Comparative Column Leaching and Pot Testing Approach.” Journal of Environmental Management 318 (September): 115530. https://doi.org/10.1016/j.jenvman.2022.115530. Chapman, Ellen Emily Vanessa, Christine Moore, and Linda M. Campbell. 2020. “Evaluation of a Nanoscale Zero-Valent Iron Amendment as a Potential Tool to Reduce Mobility, Toxicity, and Bioaccumulation of Arsenic and Mercury from Wetland Sediments.” Environmental Science and Pollution Research 27 (15): 18757–72. https://doi.org/10.1007/s11356-020-08347-6. DeSisto, Stephanie L., Heather E. Jamieson, and Michael B. Parsons. 2017. “Arsenic Mobility in Weathered Gold Mine Tailings under a Low-Organic Soil Cover.” Environmental Earth Sciences 76 (22): 773. https://doi.org/10.1007/s12665-017-7041-7. Environment and Climate Change Canada,. 2014. “Biological Test Method: Acute Lethality of Effluents to Daphnia Magna.” Guidance. Aem. May 9, 2014. https://www.canada.ca/en/environment-climate-change/services/wildlife-research-landscape-science/biological-test-method-publications/acute-lethality-effluents-daphnia-magna.html. Environment and Climate Change Canada. 2017. “Biological Test Method: Test for Survival, Growth and Reproduction in Sediment and Water Using the Freshwater Amphipod Hyalella Azteca.” RM/33. Method Development and Applications Unit Science and Technology Branch. Gaudet, Lauren. 2022. “BIoaccumulation of Contaminants in Amphibians in Historical Gold Mining Areas of Nova Scotia.” Honours in Biology, Halifax NS: Saint Mary’s University. Gil-Díaz, M., J. Alonso, E. Rodríguez-Valdés, J. R. Gallego, and M. C. Lobo. 2017. “Comparing Different Commercial Zero Valent Iron Nanoparticles to Immobilize As and Hg in Brownfield Soil.” Science of The Total Environment 584–585 (April): 1324–32. https://doi.org/10.1016/j.scitotenv.2017.02.011. Goulet, Richard R., and Patsy Thompson. 2018. “Bioaccumulation and Toxicity of Uranium, Arsenic, and Nickel to Juvenile and Adult Hyalella Azteca in Spiked Sediment Bioassays.” Environmental Toxicology and Chemistry 37 (9): 2340–49. https://doi.org/10.1002/etc.4179. Hou, Deyi, Abir Al-Tabbaa, David O’Connor, Qing Hu, Yong-Guan Zhu, Liuwei Wang, Niall Kirkwood, et al. 2023. “Sustainable Remediation and Redevelopment of Brownfield Sites.” Nature Reviews Earth & Environment 4 (4): 271–86. https://doi.org/10.1038/s43017-023-00404-1. Intrinsik Corp, EcoMetrix, Klohn Crippen Berger and Wood. 2019a. “Conceptual Closure Plan For The Historic Goldenville Tailings Areas Goldenville, Nova Scotia Final Report.” Consulting. http://www.ap.smu.ca/~lcampbel/NSLands-Historic-Goldenville-Mine-Conceptual-Closure-Plan-2019-07-25.pdf. http://nslands.ca/pdf/projects/NSLands-Historic-Goldenville-Mine-Conceptual-Closure-Plan-2019-07-25.pdf. Intrinsik Corp, EcoMetrix, Klohn Crippen Berger and Wood. 2019b. “Conceptual Closure Study For The Historic Montague Mines Tailings Areas. Halifax, Nova Scotia. Final Report.” Consulting. NS Lands Inc. http://www.ap.smu.ca/~lcampbel/NSLands-Montague-Mines-Final-Report-2019-07-24.pdf. http://nslands.ca/pdf/projects/NSLands-Montague-Mines-Final-Report-2019-07-24.pdf. Jacobs, Patrick H., and Ulrich Förstner. 1999. “Concept of Subaqueous Capping of Contaminated Sediments with Active Barrier Systems (ABS) Using Natural and Modified Zeolites.” Water Research 33 (9): 2083–87. https://doi.org/10.1016/S0043-1354(98)00432-1. Jeon, Chil-Sung, Kitae Baek, Joon-Kyu Park, Young-Ki Oh, and Sang-Do Lee. 2009. “Adsorption Characteristics of As(V) on Iron-Coated Zeolite.” Journal of Hazardous Materials 163 (2): 804–8. https://doi.org/10.1016/j.jhazmat.2008.07.052. Kang, B. Gu, and S. Park. 2016. “Evaluation of Sediment Capping Using Zeolite, Montmorillonite, and Steel Slag to Immobilize Heavy Metals.” In OCEANS 2016 - Shanghai , 1–5. https://doi.org/10.1109/OCEANSAP.2016.7485474. Knox, Anna, Michael Paller, Kenneth Dixon, Danny Reible, and Jesse Roberts. 2011. “Final Report Part II - Innovative In-Situ Remediation of Contaminated Sediments for Simultaneous Control of Contamination and Erosion.” SERDP Project ER-1501. Department of Defense Strategic Environmental Research and Development Program (SERDP). Kong, Xiangke, Zhantao Han, Wei Zhang, Le Song, and Hui Li. 2016. “Synthesis of Zeolite-Supported Microscale Zero-Valent Iron for the Removal of Cr(6+) and Cd(2+) from Aqueous Solution.” Journal of Environmental Management 169 (March): 84–90. https://doi.org/10.1016/j.jenvman.2015.12.022. LeBlanc, Molly. 2019a. “Aquatic and Emergent Invertebrate Bioaccumulation and Transfer of Mercury and Arsenic at Historical Gold Mine Tailing Sites of Nova Scotia.” Saint Mary’s University. LeBlanc, Molly E. 2019b. “Bioaccumulation and Transfer of Mercury and Arsenic in Aquatic Invertebrates and Emergent Insects at Historical Gold Mine Tailing Sites of Nova Scotia.” MSc, Halifax, Nova Scotia: Saint Mary’s University. http://library2.smu.ca/handle/01/28587. LeBlanc, Molly Evelyn, Michael B Parsons, E Emily V. Chapman, and Linda M Campbell. 2019. “Review of Ecological Mercury and Arsenic Bioaccumulation within Historical Gold Mining Districts of Nova Scotia.” Environmental Reviews , November. https://doi.org/10.1139/er-2019-0042. Lewis, Ariel S., Thomas G. Huntington, Mark C. Marvin-DiPasquale, and Aria Amirbahman. 2016. “Mercury Remediation in Wetland Sediment Using Zero-Valent Iron and Granular Activated Carbon.” Environmental Pollution 212 (May): 366–73. https://doi.org/10.1016/j.envpol.2015.11.047. Li, Juanying, Chen Zhang, Jianwei Lin, Jie Yin, Jiayan Xu, and Yiqin Chen. 2018. “Evaluating the Bioavailability of Heavy Metals in Natural-Zeolite-Amended Aquatic Sediments Using Thin-Film Diffusive Gradients.” Aquaculture and Fisheries 3 (3): 122–28. https://doi.org/10.1016/j.aaf.2018.05.003. Li, Zhangtao, Lu Wang, Jun Meng, Xingmei Liu, Jianming Xu, Fan Wang, and Philip Brookes. 2018. “Zeolite-Supported Nanoscale Zero-Valent Iron: New Findings on Simultaneous Adsorption of Cd(II), Pb(II), and As(III) in Aqueous Solution and Soil.” Journal of Hazardous Materials 344 (February): 1–11. https://doi.org/10.1016/j.jhazmat.2017.09.036. Li, Zhangtao, Lu Wang, Jizi Wu, Yan Xu, Fan Wang, Xianjin Tang, Jianming Xu, Yong Sik Ok, Jun Meng, and Xingmei Liu. 2020. “Zeolite-Supported Nanoscale Zero-Valent Iron for Immobilization of Cadmium, Lead, and Arsenic in Farmland Soils: Encapsulation Mechanisms and Indigenous Microbial Responses.” Environmental Pollution 260 (May): 114098. https://doi.org/10.1016/j.envpol.2020.114098. Liber, Karsten, Lorne E. Doig, and Suzanne L. White-Sobey. 2011. “Toxicity of Uranium, Molybdenum, Nickel, and Arsenic to Hyalella Azteca and Chironomus Dilutus in Water-Only and Spiked-Sediment Toxicity Tests.” Ecotoxicology and Environmental Safety 74 (5): 1171–79. https://doi.org/10.1016/j.ecoenv.2011.02.014. Liu, Tongzhou, Hong Wang, Zhen Zhang, and Da Zhao. 2017. “Application of Synthetic Iron-Oxide Coated Zeolite for the Pollution Control of River Sediments.” Chemosphere 180 (August): 160–68. https://doi.org/10.1016/j.chemosphere.2017.04.023. Louise Meunier, Stephen R. Walker, Joanna Wragg, Michael B. Parsons, Iris Koch, Heather E. Jamieson, and Kenneth J. Reimer. 2010. “Effects of Soil Composition and Mineralogy on the Bioaccessibility of Arsenic from Tailings and Soil in Gold Mine Districts of Nova Scotia.” Environmental Science & Technologyi 44: 2667–74. Meunier, Louise, Iris Koch, and Kenneth J. Reimer. 2011. “Effects of Organic Matter and Ageing on the Bioaccessibility of Arsenic.” Environmental Pollution , Nitrogen Deposition, Critical Loads and Biodiversity, 159 (10): 2530–36. https://doi.org/10.1016/j.envpol.2011.06.018. Mu, Yi, Zhihui Air, and Zhang, Lizhi. 2017. “Iron Oxide Shell Mediated Environmental Remediation Properties of Nano Zero-Valent Iron.” Environmental Science: Nano 4. NANOIRON. 2010. “Manual for Preparation of an Aqueous Suspension from Dry Stabilized Iron Powder NANOFER STAR.” Okamoto, Akira, Masumi Yamamuro, and Norihisa Tatarazako. 2015. “Acute Toxicity of 50 Metals to Daphnia Magna.” Journal of Applied Toxicology 35 (7): 824–30. https://doi.org/10.1002/jat.3078. Olsta, J. 2007. “In-Situ Capping of Contaminated Sediments with Reactive Materials | Ports 2007.” In . San Diego, California, S: American Society of Civil Engineers. https://ascelibrary.org/doi/abs/10.1061/40834%28238%2944. Parsons, M, K Leblanc, G Hall, A Sangster, J Vaive, and P Pelchat. 2012. “Environmental Geochemistry of Tailings, Sediments and Surface Waters Collected from 14 Historical Gold Mining Districts in Nova Scotia.” Open file 7150. Geological Survey of Canada. Saunders, Jared R., Loren D. Knopper, Iris Koch, and Kenneth J. Reimer. 2011. “Inclusion of Soil Arsenic Bioaccessibility in Ecological Risk Assessment and Comparison with Biological Effects.” Science of The Total Environment 412–413 (December): 132–37. https://doi.org/10.1016/j.scitotenv.2011.10.037. Sealy, Heather. 2011. “Arsenic Mobility and Attenuation in a Natural Wetland at Terra Mine, Northwest Territories, Canada.” Degree of Master of Science, Kingston, Ontario, Canada: Queens University. Wang, Wei, Minghua Zhou, Qiong Mao, Junjie Yue, and Xu Wang. 2010. “Novel NaY Zeolite-Supported Nanoscale Zero-Valent Iron as an Efficient Heterogeneous Fenton Catalyst.” Catalysis Communications 11 (11): 937–41. https://doi.org/10.1016/j.catcom.2010.04.004. Zhang, Chang, Meng-ying Zhu, Guang-ming Zeng, Zhi-gang Yu, Fang Cui, Zhong-zhu Yang, and Liu-qing Shen. 2016. “Active Capping Technology: A New Environmental Remediation of Contaminated Sediment.” Environmental Science and Pollution Research 23 (5): 4370–86. https://doi.org/10.1007/s11356-016-6076-8. Zhang Xin, Lin Shen, Lu XiaoQiao, and Chen ZuLiang. 2010. “Removal of Pb(II) from Water Using Synthesized Kaolin Supported Nanoscale Zero-Valent Iron.” Chemical Engineering Journal 163 (3): 243–48. https://doi.org/10.1016/j.cej.2010.07.056. Cite Share Download PDF Status: Posted Version 1 posted You are reading this latest preprint version Research Square lets you share your work early, gain feedback from the community, and start making changes to your manuscript prior to peer review in a journal. As a division of Research Square Company, we’re committed to making research communication faster, fairer, and more useful. We do this by developing innovative software and high quality services for the global research community. Our growing team is made up of researchers and industry professionals working together to solve the most critical problems facing scientific publishing. Also discoverable on Platform About Our Team In Review Editorial Policies Advisory Board Help Center Resources Author Services Accessibility API Access RSS feed Manage Cookie Preferences © Research Square 2026 | ISSN 2693-5015 (online) Privacy Policy Terms of Service Do Not Sell My Personal Information {"props":{"pageProps":{"initialData":{"identity":"rs-3894488","acceptedTermsAndConditions":true,"allowDirectSubmit":true,"archivedVersions":[],"articleType":"Research Article","associatedPublications":[],"authors":[{"id":282003209,"identity":"372a903a-be97-47c4-8f9d-7e19b56927af","order_by":0,"name":"Ellen Emily Vanessa Chapman","email":"data:image/png;base64,iVBORw0KGgoAAAANSUhEUgAAAZAAAAAyAQMAAABI0h/eAAAABlBMVEX///8AAABVwtN+AAAACXBIWXMAAA7EAAAOxAGVKw4bAAAA5ElEQVRIiWNgGAWjYBAC+wbGxgMMDHYJEG4FEVoMGBgbgFqSoVrOEKWFgQGo5SBEC2MbMVrEDjcc5mE4kMc/uzvxc+G8umiDA8wPP+DTYi+dCNZSLHHn7GbpmdsO5244wGYsgdcWqJbEhhu5G6R5tx0AauFhIE7L/Bu5m3/zzqkDaWH+QYSWg4kbbuRuk+ZtYAZpYSNoy8E5BsmJG++c3WbNc+xw7szDbGYW+LTYz05/+OBNhV3ivNu9m2/z1NTl9h1vfnwDnxYQYOIBxQ7cMcyE1AMBI9i3eN0/CkbBKBgFIxoAACrfU30M9pGrAAAAAElFTkSuQmCC","orcid":"https://orcid.org/0000-0002-3654-9113","institution":"Saint Mary's University","correspondingAuthor":true,"prefix":"","firstName":"Ellen","middleName":"Emily Vanessa","lastName":"Chapman","suffix":""},{"id":282003210,"identity":"72b9a628-5380-4d6f-8079-45387ab63d27","order_by":1,"name":"Linda M. Campbell","email":"","orcid":"","institution":"Saint Mary's University","correspondingAuthor":false,"prefix":"","firstName":"Linda","middleName":"M.","lastName":"Campbell","suffix":""}],"badges":[],"createdAt":"2024-01-24 15:20:46","currentVersionCode":1,"declarations":"","doi":"10.21203/rs.3.rs-3894488/v1","doiUrl":"https://doi.org/10.21203/rs.3.rs-3894488/v1","draftVersion":[],"editorialEvents":[],"editorialNote":"","failedWorkflow":false,"files":[{"id":53384511,"identity":"03316803-95a3-4777-b4a5-1439c8d21041","added_by":"auto","created_at":"2024-03-25 10:57:36","extension":"jpg","order_by":1,"title":"Figure 1","display":"","copyAsset":false,"role":"figure","size":260609,"visible":true,"origin":"","legend":"\u003cp\u003eSummary and timeline of the experimental design with treatments, replicates and tests. a) On day 0, the three different types of sediment were added to 18 1-L beakers each and samples were collected of sediments from the 54 beakers for meta(oid) analysis. Following sediment sampling, the 18 beakers representing each sediment type were divided into three groups of six, which either received no treatment, treatment “A” or “B” as outlined in the experimental design section. \u003cem\u003eHyalella azteca\u003c/em\u003e culture water was then added to each of the 54 beakers and were allowed to settle for 10 days. b) On day 10, the 5-day Hg bioaccumulation test with amano shrimp (\u003cem\u003eCaridina multidentate\u003c/em\u003e) and spike-topped apple snails (\u003cem\u003ePomacea bridgesi\u003c/em\u003e) was started. Water samples were collected for metal(oid) analysis at the end of this test (day 15). c) The 14-day \u003cem\u003eHyalella azteca\u003c/em\u003etoxicity test was initiated on day 20. Water samples and porewater samples were collected for metal(oid) analysis at the end of this test (day 34). d) The 48 hour \u003cem\u003eDaphnia magna\u003c/em\u003e toxicity test was started on day 34 and finalized on day 36\u003c/p\u003e","description":"","filename":"Fig1.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/fc6977fb327da8c2a8003cc4.jpg"},{"id":53383823,"identity":"8680755a-53d9-4ca3-b54f-8aaaa43948eb","added_by":"auto","created_at":"2024-03-25 10:49:36","extension":"jpg","order_by":2,"title":"Figure 2","display":"","copyAsset":false,"role":"figure","size":180652,"visible":true,"origin":"","legend":"\u003cp\u003eScanning Electron Microscope (SEM) images and atomic weight composition (%) a) dry zeolite, b) nZVI slurry and c) zeolite-n-ZVI slurry blend. The nZVI particles in (b) are showing extensive agglomeration. When nZVI blended with zeolite (c), nZVI particles are still showing agglomeration, but not to the same extent as in (b) suggesting that a simple mechanical blend of these two materials can increase surface area and reactivity of nZVI. Atomic weight composition suggests changes in nZVI distribution across the surface of zeolite. nZVI particles appear concentrated in cracks or pores on the zeolite particles surface\u003c/p\u003e","description":"","filename":"Fig2.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/2c66fff0c11380e46c53bcbd.jpg"},{"id":53383822,"identity":"85e9cb95-59c3-4c1e-b762-108b18d877ce","added_by":"auto","created_at":"2024-03-25 10:49:36","extension":"jpg","order_by":3,"title":"Figure 3","display":"","copyAsset":false,"role":"figure","size":267929,"visible":true,"origin":"","legend":"\u003cp\u003ePorewater concentrations of dissolved Hg in untreated sediment (black circles), sediment with treatment A (grey circles), and sediment with treatment B (white circles). C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig3.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/18b80a5d78a6b3babbfa9465.jpg"},{"id":53383820,"identity":"bd2a990a-6552-4753-a8c5-efc8abab5437","added_by":"auto","created_at":"2024-03-25 10:49:36","extension":"jpg","order_by":4,"title":"Figure 4","display":"","copyAsset":false,"role":"figure","size":538216,"visible":true,"origin":"","legend":"\u003cp\u003eConcentrations of As (Total and Dissolved) in overlaying water column of treated and untreated sediments on day 15 (a) and 34 (b). Dotted line represents CCME guideline value for protection of aquatic life (5 µg/L). C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig4.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/a834a56915b0b90d42e4411d.jpg"},{"id":53383825,"identity":"4ac28bed-01a5-4c30-bcaa-5e17aa53fbab","added_by":"auto","created_at":"2024-03-25 10:49:36","extension":"jpg","order_by":5,"title":"Figure 5","display":"","copyAsset":false,"role":"figure","size":456803,"visible":true,"origin":"","legend":"\u003cp\u003eConcentrations of Hg (Total and Dissolved) in overlaying water column of treated and untreated sediments on day 15 (a) and 34 (b). C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig5.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/1f972b28571fa6350282a806.jpg"},{"id":53384512,"identity":"a1f7bcaf-1a0c-40a1-8179-d6d9cedfc087","added_by":"auto","created_at":"2024-03-25 10:57:36","extension":"jpg","order_by":6,"title":"Figure 6","display":"","copyAsset":false,"role":"figure","size":218274,"visible":true,"origin":"","legend":"\u003cp\u003eShrimp (\u003cem\u003eCaridina multidentate\u003c/em\u003e) tissue concentrations of Hg in after 5 day exposure to untreated sediment (black boxes), sediment with treatment A (grey boxes), and sediment with treatment B (white boxes). C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig6.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/dd131953aed846a0752dfd78.jpg"},{"id":53383826,"identity":"4d394d73-dfb8-4a56-9e62-d1d7dc4a160b","added_by":"auto","created_at":"2024-03-25 10:49:37","extension":"jpg","order_by":7,"title":"Figure 7","display":"","copyAsset":false,"role":"figure","size":281802,"visible":true,"origin":"","legend":"\u003cp\u003e\u003cem\u003eHyalella azteca\u003c/em\u003e survival (%) after 14 days exposure to treated and untreated sediments. C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig7.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/3b6899f6888bd2c1b36cbcbf.jpg"},{"id":53383827,"identity":"76c9b8cc-c0b9-4c60-a649-34dcbad0f9c5","added_by":"auto","created_at":"2024-03-25 10:49:37","extension":"jpg","order_by":8,"title":"Figure 8","display":"","copyAsset":false,"role":"figure","size":278562,"visible":true,"origin":"","legend":"\u003cp\u003e\u003cem\u003eDaphnia magna\u003c/em\u003e survival (%) of after 48 hours exposure to treated and untreated sediments. C=control sediment, MP=Muddy Pond sediment, OS=Old Stamp Mill Sediment, A=treatment A, B=treatment B\u003c/p\u003e","description":"","filename":"Fig8.jpg","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/8f82d58574068b1d0c3c6cc5.jpg"},{"id":54729467,"identity":"c29a24f2-3215-44dd-8c85-6883b89394f1","added_by":"auto","created_at":"2024-04-15 21:24:41","extension":"pdf","order_by":0,"title":"","display":"","copyAsset":false,"role":"manuscript-pdf","size":1140307,"visible":true,"origin":"","legend":"","description":"","filename":"manuscript.pdf","url":"https://assets-eu.researchsquare.com/files/rs-3894488/v1/7de268bf-b24f-41b1-a012-812d60a8c4b8.pdf"}],"financialInterests":"","formattedTitle":"Evaluation of a customized reactive nanoscale-zero-valent iron and zeolite thin capping blend for enhancing natural recovery of wetlands impacted by contaminated legacy gold mine tailings","fulltext":[{"header":"Introduction","content":"\u003cp\u003eWith 360 gold mines in 64 historic gold mining districts across Nova Scotia (NS) Canada, there is a significant legacy of potentially toxic mining waste \u0026ldquo;tailings\u0026rdquo; in this province. It has been estimated that over 3\u0026nbsp;million tonnes of tailings were generated during the historical gold rushes in NS between the 1860\u0026rsquo;s \u0026amp; 1940\u0026rsquo;s (Parsons et al. \u003cspan citationid=\"CR37\" class=\"CitationRef\"\u003e2012\u003c/span\u003e). Tailings contain elevated concentrations of mercury (Hg) due to the mercury amalgamation process used to extract the gold, and arsenic (As) due to the geogenic arsenopyrite in gold-bearing rock (Meunier et al., 2010). Gold amalgamation processes in the late 1800\u0026rsquo;s required freshwater to process crushed ore, so legacy gold mine ore processing and tailing sites are generally situated close to low-lying lakes, rivers, streams or wetlands. After ore processing, the tailing material was often slurried back into the same freshwater bodies without treatment. Consequently, aquatic ecosystems in historical gold mine districts are at particular risk, even over 100 years later. Despite this, impacts of tailings on aquatic ecosystems and freshwater biota from legacy gold mines remain poorly understood ( LeBlanc et al. \u003cspan citationid=\"CR24\" class=\"CitationRef\"\u003e2019\u003c/span\u003e). Multi-generational chronic exposure to toxic tailings has occurred in many wetlands and lakes, potentially impairing ecosystem function and impeding biological recovery(Alpers et al. \u003cspan citationid=\"CR1\" class=\"CitationRef\"\u003e2005\u003c/span\u003e; Alpers et al. \u003cspan citationid=\"CR2\" class=\"CitationRef\"\u003e2016\u003c/span\u003e; AQUAMIN Steering Group \u003cspan citationid=\"CR3\" class=\"CitationRef\"\u003e1996\u003c/span\u003e). Over the past three years, our group has made significant headway on assessing the impacts of contaminated tailings to freshwater ecosystems in the field and in the laboratory (Chapman, et al. \u003cspan citationid=\"CR7\" class=\"CitationRef\"\u003e2020\u003c/span\u003e; LeBlanc \u003cspan citationid=\"CR22\" class=\"CitationRef\"\u003e2019a\u003c/span\u003e; LeBlanc \u003cspan citationid=\"CR23\" class=\"CitationRef\"\u003e2019b\u003c/span\u003e; Gaudet \u003cspan citationid=\"CR11\" class=\"CitationRef\"\u003e2022\u003c/span\u003e). We have shown that sediment and water at these sites are acutely toxic to sensitive aquatic invertebrates, and that more tolerant aquatic invertebrates and amphibians in impacted wetlands bioaccumulate significantly higher Hg and As concentrations than those from reference wetlands. We also examined As and Hg concentrations in adult dragonflies and their juvenile aquatic nymph counterparts, and have shown that the aquatic nymphs accumulate significant As and Hg concentrations underwater (LeBlanc \u003cspan citationid=\"CR23\" class=\"CitationRef\"\u003e2019b\u003c/span\u003e). When the nymphs emerge as adults, the adult dragonflies tend to have consistently elevated Hg concentrations and will have elevated As concentrations at sites with high levels of As in sediments. Those invertebrates present an ecosystem risk because top predators, both underwater (fish, diving birds) and terrestrial (bats, birds, insect-eating mammals) will be exposed to significant Hg and As through consuming those. Because Hg and As can be transferred up food webs, economically-valuable species and species at risk likely are impacted which complicate future environmental assessments and raises liability issues for Crown land. Natural recovery strategies (i.e., leaving these sites alone) have clearly not been successful, with many legacy gold-mine-tailing wetlands still exhibiting severe contamination issues more than 100 years after the historical gold rushes ended.\u003c/p\u003e \u003cp\u003eNova Scotia wetland and shallow-water sites impacted by legacy gold-mine tailing material are frequently situated in remote areas and are near high-value freshwater ecosystems supporting economically valuable fisheries and recreation sites. Dredging to remove contaminated material would be prohibitive at this provincial scale, and can cause a re-release of As and Hg back into freshwater ecosystems (DeSisto et al. \u003cspan citationid=\"CR8\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Large remediation projects involving excavation and off-site treatment would also consume significant amounts of energy and emit large quantities of greenhouse gases (Hou et al. \u003cspan citationid=\"CR14\" class=\"CitationRef\"\u003e2023\u003c/span\u003e). The government of Nova Scotia recently estimated the cost of remediation of the Crown land portion of two legacy gold mining sites at \u003cspan\u003e$\u003c/span\u003e60\u0026nbsp;million CAD (Intrinsik Corp, et al. 2019a, 2019b). High-level conceptual closure plans prepared for these two sites included recommendations to move contaminated tailings on land to containment cells, while contaminated wetland areas be capped due to the risk associated with disturbing wetland sediments.\u003c/p\u003e \u003cp\u003eCapping of freshwater sediments can provide cost-effective isolation and exposure-pathway elimination. Conventional (passive) caps or low permeability liners, normally uses clean, neutral materials like sand, silt, clay, and crushed-rock debris and rely on containment rather than treatment (Zhang et al. \u003cspan citationid=\"CR41\" class=\"CitationRef\"\u003e2016\u003c/span\u003e). These materials are often easy to find at a moderate cost. However, these types of caps require a thick layer of material, often at least 50 cm to be effective in physical isolation of contaminated sediment. In shallow wetland environments commonly associated with Nova Scotia sites, reductions in wetland flood storage capacity and depth of the overlaying water needs to be considered. There is a need for innovative, resource-efficient in-situ remediation strategies that can cost-effectively manage the risk of both As and Hg in wetland sediments while retaining and improving wetland functionality (Intrinsik Corp, et al. 2019a, 2019b).\u003c/p\u003e \u003cp\u003eThe thickness, function, and longevity of traditional, passive isolation caps can be improved through the use of reactive capping materials, which ideally reduce contaminants\u0026rsquo; mobility, toxicity, and bioavailability and thereby offer both containment and treatment of contaminated sediment. Olsta (\u003cspan citationid=\"CR36\" class=\"CitationRef\"\u003e2007\u003c/span\u003e) states that a 12 mm (0.5 in) thick reactive mat can theoretically replace 1 m (3 ft) of sand or soil. Reactive materials that have been used in cappings of contaminated sediments include activated carbon, biochar, composts, organoclays, calcite, zeolite, bentonite, apatite, biopolymers, and zerovalent iron (ZVI) (Zhang et al. \u003cspan citationid=\"CR41\" class=\"CitationRef\"\u003e2016\u003c/span\u003e). Unfortunately, common reactive soil amendments, especially organic amendments used to reduce bioavailability of some metals can significantly increase the mobility and bioaccessibility of As in wet soils and sediment, leading to elevated human and ecological risks (Cerqueira et al. \u003cspan citationid=\"CR6\" class=\"CitationRef\"\u003e2022\u003c/span\u003e; Meunier et al. 2011; Saunders et al. \u003cspan citationid=\"CR38\" class=\"CitationRef\"\u003e2011\u003c/span\u003e). Consequently, using organic amendments and fertilizer additions, which is common practice for phytostabilization projects elsewhere, may not be a feasible stand-alone option for historical gold-mining tailing sites with elevated As and fluctuating water levels (Meunier et al. 2011). Knox et al., (\u003cspan citationid=\"CR20\" class=\"CitationRef\"\u003e2011\u003c/span\u003e) studied multiple amendment active caps (MAACs) for the remediation of contaminated sediments, which consist of a mixture of chemically active amendments combined with sand or other neutral materials. They found that phosphate, zeolite, bentonite, and organoclays individually or mixed with another active or neutral materials can stabilize metals and nonpolar pollutants (e.g., PAHs), and addition of a small amount of bentonite (e.g., 10%) to MAACs can improve erosion resistance and metal sequestration capacity. Zeolites are crystalline hydrated aluminosilicates of alkali and alkaline earth elements with a very high cation exchange capacity of up to 6 mmol(eq)/g. Due to the high cation exchange capacity, natural zeolites are capable of demobilising large amounts of cationic pollutants by sorption (Jacobs and F\u0026ouml;rstner 1999), but zeolite has a very low affinity for anionic compounds because of negative surface charges. Jeon et al., (\u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2009\u003c/span\u003e) tested iron (III) coated zeolite for As(V) removal and found that As(V) was combined with iron oxyhydroxide onto the zeolite by complexation, but the adsorption capacity was insufficient to ameliorate high contamination levels.\u003c/p\u003e \u003cp\u003eLi et al., (\u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e) suggested that nanoscale zero-valent iron (nZVI) with a high anionic adsorption capacity and unique core-shell structures (Mu et al. \u003cspan citationid=\"CR33\" class=\"CitationRef\"\u003e2017\u003c/span\u003e), combined with zeolite may overcome the deficiencies of zeolite for treatment of As. nZVI is also relatively low-cost, in comparison with other metallic nanoparticles (Arshadi et al. \u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2017\u003c/span\u003e; Gil-D\u0026iacute;az et al. \u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2017\u003c/span\u003e), and is an effective reducing agent for metal (loid)s (Arshadi et al. \u003cspan citationid=\"CR5\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). ZVI treatment has been shown to lower the toxicity of As and Hg through converting oxidized elements to less mobile and toxic forms. Additionally, all nZVI particles consists of a zero-valent iron (Fe0) core and an iron oxide shell structure. The nature of the iron oxide shell on the Fe0 core of nZVI particles has an important impact on the inorganic contaminant removal performance of nZVI. It has been show that the adsorption of inorganic species by nZVI was predominantly mediated by this oxide shell (Mu et al. \u003cspan citationid=\"CR33\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Mu et al. (\u003cspan citationid=\"CR33\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) reported that the oxide shell might consist of a single phase or be made up of several phases (such as w\u0026uuml;stite (FeO), magnetite (Fe3O4), maghemite (γ-Fe2O3), hematite (α-Fe2O3), and goethite (FeOOH)).\u003c/p\u003e \u003cp\u003eIn an earlier study: (Chapman et al. \u003cspan citationid=\"CR7\" class=\"CitationRef\"\u003e2020\u003c/span\u003e), we explored whether nZVI added to two different contaminated wetland sediments could reduce Hg and As mobility and toxicity to two aquatic invertebrates; burrowing mayflies (\u003cem\u003eHexagenia spp\u003c/em\u003e) and Chinese mystery snails (\u003cem\u003eCipangopaludina chinensis\u003c/em\u003e). Total Hg and As concentrations in overlaying water above both contaminated sediments were reduced by at least 75% and 88% respectively when treated with nZVI. In the first sediment, juvenile snail survival increased from 75% in the untreated sediment to 100% in all nZVI treatments. The 2% nZVI treatment level was the only one with surviving mayflies (33%) and growth of juvenile snails. No snails or mayflies survived in the second sediment with higher Hg and As concentrations regardless of nZVI treatment level. It is possible that this is because nZVI tends to aggregate rapidly in water, which reduces its adsorption capacity significantly. In order to prevent this aggregation of nZVI, different types of clay minerals have been successfully used to support nZVI, such as zeolite (Kong et al. \u003cspan citationid=\"CR21\" class=\"CitationRef\"\u003e2016\u003c/span\u003e; Li et al. \u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e; Wang et al. \u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2010\u003c/span\u003e).\u003c/p\u003e \u003cp\u003eZeolite-supported nZVI has mostly been synthesized in laboratories using liquid phase reduction and ion exchange procedures (Kong et al. \u003cspan citationid=\"CR21\" class=\"CitationRef\"\u003e2016\u003c/span\u003e; Li et al. \u003cspan citationid=\"CR28\" class=\"CitationRef\"\u003e2020\u003c/span\u003e; \u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e; Wang et al. \u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2010\u003c/span\u003e). In short, this involves first blending zeolite with ferric chloride or FeSO\u003csub\u003e4\u003c/sub\u003e7H\u003csub\u003e2\u003c/sub\u003eO in different ratios with water, pH-adjusted and stirred vigorously. Next, aNaBH\u003csub\u003e4\u003c/sub\u003e solution is added to ensure adequate reduction of Fe(III). Finally, the zeolite-nZVI composite is separated from the mixture solution by magnet, centrifuge, or vacuum filtration, washed with ultrapure water to remove soluble impurities, and then usually stored in vacuum before use. This process is time consuming and requires laboratory-grade equipment and chemicals in large quantities, making this impractical for large-scale field application. Commercially-available sources of nZVI and zeolite and simpler production methods for blends of these ingredients would be preferable for bulk distribution in wetland areas. For example, Liu et al., (\u003cspan citationid=\"CR30\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) tested an iron-oxide coated zeolite for the pollution control of river sediments. The iron-oxide coated zeolite in this study was prepared by a simple mixing of moistened iron oxide powder with deionized water and zeolite. This iron-oxide coated product showed capacities to sorb ammonia, phosphates and sulfides from sediments. It may therefore be possible to coat zeolite with nZVI particles in a similar procedure as outlined for iron-oxides by Liu et al. (\u003cspan citationid=\"CR30\" class=\"CitationRef\"\u003e2017\u003c/span\u003e), but this has not yet been investigated.\u003c/p\u003e \u003cp\u003eNANOFER STAR is a commercially available prepared dry air-stable nZVI powder, where the surface of iron nanoparticles is stabilized by a thin layer of iron oxide, which prevents immediate oxidation in contact with atmospheric oxygen. The nanoparticles of this product are in form of clusters and agglomerates and it is therefore necessary to \u0026ldquo;activate\u0026rdquo; the nanoparticles by dispersing them and eroding the iron oxide layer in an aqueous suspension (slurry). According to the manufacturer, the highest reactivity of this slurry was following 48 hours after suspension of powder in water (NANOIRON \u003cspan citationid=\"CR34\" class=\"CitationRef\"\u003e2010\u003c/span\u003e). NANOFER STAR product becomes very reactive in water environment, hydrogen gas is produced during reaction of the product with water and Fe(0) nanoparticles are transformed to iron oxides and hydroxides. Therefore, it is necessary to apply the slurry as soon as possible following the 48-hour activation period. Zeolite has not been added to this product, but if incorporated during the mixing and activation period, we hypothesize that nZVI will coat the zeolite, which in turn could lead to reduced agglomeration and enhanced reactivity for Hg and As contamination.\u003c/p\u003e \u003cp\u003eThe objectives of this study were: 1) to test a simple method for coating the NANOFER STAR nZVI on zeolite, and 2) to test this nZVI coated zeolite blend as a reactive barrier under a protective capping layer of sand and clays, and as a component of the protective capping in an \u0026ldquo;active cap\u0026rdquo; for limiting the mobility, bioaccumulation and toxicity of Hg and As in wetland sediment impacted by legacy gold mine tailings.\u003c/p\u003e"},{"header":"Methods","content":"\u003cdiv id=\"Sec3\" class=\"Section2\"\u003e \u003ch2\u003eContaminated sediment collection and control sediment preparation\u003c/h2\u003e \u003cp\u003eThe contaminated wetland sites for this study were the Muddy Pond (MP) wetland, located within the Waverley historical gold mining district and the Old Stamp Mill (OS) wetland, located within the Montague historical gold mining district, both in the Halifax Regional Municipality, Nova Scotia, Canada. We collected 18 10-cm deep soil cores (diameter 7\u0026ndash;8 cm) from each site, which were immediately inserted into heavy-duty zip-lock bags. Care was taken to ensure minimal disturbance of the sediment and contact with air. The sediment was stored cold (7˚C), in vacuum, and in the dark until the start of the test.\u003c/p\u003e \u003cp\u003eAn artificial control sediment was prepared with the goal to mimic the texture and pH of the tailing sediments. This consisted of 90% fine sand (\u0026lt;\u0026thinsp;180 um), 5% peat (sieved to \u0026lt;\u0026thinsp;1mm) and 5% clay. CaCO\u003csub\u003e3\u003c/sub\u003e was added to the sediment in order to increase pH to 6\u0026ndash;7.\u003c/p\u003e \u003cp\u003eParticle size distribution of the wetland and control sediments was determined using laser diffraction particle size analysis by Loring Tarcore Labs Ltd for clay content (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e). Organic matter was determined by percent loss on ignition (% LOI). Dry sediment material was weighed and placed in a muffle furnace at 550\u0026deg;C for exactly 4 h. Samples were then re-weighed and % loss on ignition was calculated from sediment weight differences between 60˚C and 550\u0026deg;C (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e).\u003c/p\u003e \u003cp\u003e \u003cdiv class=\"gridtable\"\u003e\u003ctable float=\"Yes\" id=\"Tab1\" border=\"1\"\u003e \u003ccaption language=\"En\"\u003e \u003cdiv class=\"CaptionNumber\"\u003eTable 1\u003c/div\u003e \u003cdiv class=\"CaptionContent\"\u003e \u003cp\u003eSediment type and properties (metal(oid)s, sulphur, organic matter content as determined by loss on ignition (LOI), clay content, and pH over overlaying water at test start)\u003c/p\u003e \u003c/div\u003e \u003c/caption\u003e \u003ccolgroup cols=\"5\"\u003e \u003cdiv align=\"left\" class=\"colspec\" colname=\"c1\" colnum=\"1\"\u003e\u003c/div\u003e \u003cdiv align=\"left\" class=\"colspec\" colname=\"c2\" colnum=\"2\"\u003e\u003c/div\u003e \u003cdiv align=\"left\" class=\"colspec\" colname=\"c3\" colnum=\"3\"\u003e\u003c/div\u003e \u003cdiv align=\"left\" class=\"colspec\" colname=\"c4\" colnum=\"4\"\u003e\u003c/div\u003e \u003cdiv align=\"left\" class=\"colspec\" colname=\"c5\" colnum=\"5\"\u003e\u003c/div\u003e \u003cthead\u003e \u003ctr\u003e \u003cth align=\"left\" colname=\"c1\"\u003e \u003cp\u003eSediment type/properties\u003c/p\u003e \u003c/th\u003e \u003cth align=\"left\" colname=\"c2\"\u003e \u003cp\u003eMuddy Pond (MP)\u003c/p\u003e \u003c/th\u003e \u003cth align=\"left\" colname=\"c3\"\u003e \u003cp\u003eOld Stamp Mill (OS)\u003c/p\u003e \u003c/th\u003e \u003cth align=\"left\" colname=\"c4\"\u003e \u003cp\u003eControl/artificial sediment (C)\u003c/p\u003e \u003c/th\u003e \u003cth align=\"left\" colname=\"c5\"\u003e \u003cp\u003eCCME PEL (mg/kg)\u003c/p\u003e \u003c/th\u003e \u003c/tr\u003e \u003c/thead\u003e \u003ctbody\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eAs (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e94 211\u0026thinsp;\u0026plusmn;\u0026thinsp;25 437\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e788\u0026thinsp;\u0026plusmn;\u0026thinsp;560\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e0.3\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e17\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eHg (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e41\u0026thinsp;\u0026plusmn;\u0026thinsp;13\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e91\u0026thinsp;\u0026plusmn;\u0026thinsp;16\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e0.012\u0026thinsp;\u0026plusmn;\u0026thinsp;0.004\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e0.486\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003ePb (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e716\u0026thinsp;\u0026plusmn;\u0026thinsp;224\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e710\u0026thinsp;\u0026plusmn;\u0026thinsp;112\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e5.0\u0026thinsp;\u003cb\u003e\u0026plusmn;\u003c/b\u003e\u0026thinsp;0.3\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e91.3\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eCu (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e54\u0026thinsp;\u0026plusmn;\u0026thinsp;15\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e478\u0026thinsp;\u0026plusmn;\u0026thinsp;71\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e0.7\u0026thinsp;\u003cb\u003e\u0026plusmn;\u003c/b\u003e\u0026thinsp;0.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e197\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eNi (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e133\u0026thinsp;\u0026plusmn;\u0026thinsp;38\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e18\u0026thinsp;\u0026plusmn;\u0026thinsp;2\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e2.0\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e75\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eFe (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e95 161\u0026thinsp;\u0026plusmn;\u0026thinsp;24 823\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e19 728\u0026thinsp;\u0026plusmn;\u0026thinsp;1513\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e489\u0026thinsp;\u0026plusmn;\u0026thinsp;47\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e43 766\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eAg (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e1.4\u0026thinsp;\u0026plusmn;\u0026thinsp;0.6\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e2.0\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e0.008\u0026thinsp;\u003cb\u003e\u0026plusmn;\u003c/b\u003e\u0026thinsp;0.001\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e1\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eZn (mg/kg) (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e\u003cb\u003e657\u0026thinsp;\u0026plusmn;\u0026thinsp;140\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e\u003cb\u003e314\u0026thinsp;\u0026plusmn;\u0026thinsp;34\u003c/b\u003e\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e1.3\u0026thinsp;\u003cb\u003e\u0026plusmn;\u003c/b\u003e\u0026thinsp;0.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e315\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eS% (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e4.7\u0026thinsp;\u0026plusmn;\u0026thinsp;1.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e0.03\u0026thinsp;\u0026plusmn;\u0026thinsp;0.02\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e0.02\u0026thinsp;\u0026plusmn;\u0026thinsp;0\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e-\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003eLOI% (n\u0026thinsp;=\u0026thinsp;18)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e9.3\u0026thinsp;\u0026plusmn;\u0026thinsp;2.9\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e3.8\u0026thinsp;\u0026plusmn;\u0026thinsp;1.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e6.4\u0026thinsp;\u0026plusmn;\u0026thinsp;0.6\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e-\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003e% Clay (n\u0026thinsp;=\u0026thinsp;1)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e4.4\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e4.5\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e5.0\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e-\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003ctr\u003e \u003ctd align=\"left\" colname=\"c1\"\u003e \u003cp\u003epH (overlaying water) (n\u0026thinsp;=\u0026thinsp;6)\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c2\"\u003e \u003cp\u003e6.9\u0026thinsp;\u0026plusmn;\u0026thinsp;0.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c3\"\u003e \u003cp\u003e6.4\u0026thinsp;\u0026plusmn;\u0026thinsp;0.1\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c4\"\u003e \u003cp\u003e7.5\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2\u003c/p\u003e \u003c/td\u003e \u003ctd align=\"left\" colname=\"c5\"\u003e \u003cp\u003e-\u003c/p\u003e \u003c/td\u003e \u003c/tr\u003e \u003c/tbody\u003e \u003c/colgroup\u003e \u003c/table\u003e\u003c/div\u003e \u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec4\" class=\"Section2\"\u003e \u003ch2\u003enZVI/zeolite reactive blend and protective capping material preparation\u003c/h2\u003e \u003cp\u003eA NANOFERSTAR nZVI (NANOIRON, \u003cspan citationid=\"CR34\" class=\"CitationRef\"\u003e2010\u003c/span\u003e) stock slurry was prepared by blending 50 g of nZVI (with ~\u0026thinsp;80% active Fe\u003csup\u003e0\u003c/sup\u003e ) and 200 mL of dechlorinated water for 10 minutes in 1000W blender (NANOIRON \u003cspan citationid=\"CR34\" class=\"CitationRef\"\u003e2010\u003c/span\u003e). Next, 80g of Zeolite (ultrafine\u0026thinsp;\u0026lt;\u0026thinsp;20\u0026micro;m, 90\u0026ndash;92% Clinoptilolite from Heiltropfen\u0026reg;) was added to the nZVI stock slurry and blended for another 10 minutes aiming for a ratio of 2:1 zeolite/nZVI as per Arancibia-Miranda et al., (\u003cspan citationid=\"CR4\" class=\"CitationRef\"\u003e2016\u003c/span\u003e) and Wang et al., (\u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2010\u003c/span\u003e). The Zeolite/nZVI slurry was then activated by storing for 48 hrs in a sealed plastic amber PET bottle at room temperature. It was observed that following the activation period, the hydrogen production (visually assessed by bulging of the PET bottle) was much greater in the zeolite/nZVI slurry compared with a control slurry containing nZVI only, which indicates higher reactivity in the zeolite/nZVI slurry.\u003c/p\u003e \u003cp\u003eThe protective capping layer was prepared by mixing fine silica sand (80%), bentonite - montmorillonite (10%) and ultrafine micronized (\u0026lt;\u0026thinsp;20um) zeolite -clinoptilolite (10%).\u003c/p\u003e \u003cp\u003eSamples of NANOFERSTAR nZVI activated slurry, zeolite powder, and nZVI-coated zeolite activated slurry were cold mounted in a polished epoxy resin. The surface morphology, particle size and elemental distributions were then determined using the TESCAN Mira3 LMU scanning electronic microscope (SEM) (maximum resolution up to 1.2 nm at 30 kV combined with an energy dispersive X-ray (EDX) analysis) in the Saint Mary\u0026rsquo;s University Electron Microscopy Centre.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec5\" class=\"Section2\"\u003e \u003ch2\u003eExperimental design\u003c/h2\u003e \u003cp\u003eThe multi-phase experimental design (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) included a total of 54 1-L beakers, with 18 beakers containing 100-mL of each sediment type (MP, OS and control). Within those 18 beakers, 6 replicates were left untreated while the other 12 received two different treatments: \u0026ldquo;A\u0026rdquo; or \u0026ldquo;B\u0026rdquo; (each with 6 replicates) as outlined below.\u003c/p\u003e \u003cp\u003eTreatment \u0026ldquo;A\u0026rdquo; consisted of 5% nZVI coated zeolite slurry added to the top of each sediment based on wet weight of sediment, then 50 ml (1cm) protective capping material was added on top of this. The 5% nZVI coated zeolite concentration equates to 30 g nZVI coated zeolite per kg of sediment. This concentration was selected as Li et al. (\u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e) found that arsenic, cadmium and lead in soil samples were immobilized at this concentration with their zeolite-supported nZVI. Treatment \u0026ldquo;B\u0026rdquo; consisted of a 100 ml (2-cm) layer of protective capping material mixed with the nZVI coated zeolite at 5% of wet weight of capping added to the top of each sediment.\u003c/p\u003e \u003cp\u003eAll beakers then received 650 ml (9-cm water column above the top of the sediment) of specially prepared \u0026ldquo;\u003cem\u003eHyalella azteca\u003c/em\u003e culture water\u0026rdquo;. This culture water was prepared based on federal government ecotoxicology standard protocols (Environment and Climate Change Canada \u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Beakers with sediment and water were then allowed to settle for 10 days before starting the Hg bioaccumulation experiment.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec6\" class=\"Section2\"\u003e \u003ch2\u003eHg bioaccumulation experiment\u003c/h2\u003e \u003cp\u003eLarge-bodied invertebrates used for bioaccumulation studies (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) included Amano shrimp (\u003cem\u003eCaridina multidentate\u003c/em\u003e) and spike-topped apple snails (\u003cem\u003ePomacea bridgesi\u003c/em\u003e). These were selected for their relative tolerance of contamination and large individual biomass, allowing tissue analysis of Hg. Three replicates of each treatment #1\u0026ndash;3 received one snail each, and three replicates of each treatment #4\u0026ndash;6 received one shrimp. Snails and shrimp were left unfed for 5 days; the duration of the test, which was determined by a preliminary experiment showing that shrimp could not survive past 7 days in jars with clean silica sand without feeding. Following the 5-day exposure period, the shrimp and snails were retrieved and allowed to depurate in clean test water for 24 hours. Snails and shrimp were euthanized by rapidly freezing them at -20C˚. All invertebrates were then individually weighed, and dried at \u0026le;\u0026thinsp;60℃, then weighed again to determine moisture content. After removing snails from their shells, dried whole-body tissues were ground to a powder using an Retsch MM 400 ball mill mixer prior to analysis.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec7\" class=\"Section2\"\u003e \u003ch2\u003eInvertebrate ecotoxicity experiments\u003c/h2\u003e \u003cp\u003eFive days after the snails and shrimp were removed from the beakers, a 14-day toxicity test with \u003cem\u003eHyalella azteca\u003c/em\u003e was initiated (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) in the same beakers, following Environment and Climate Change Canada (\u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) ecotoxicology protocols. Briefly, ten juvenile \u003cem\u003eHyalella\u003c/em\u003e (age 1\u0026ndash;7 days) were placed in each of 5 replicates/treatment for a 14-day exposure. Lights were set to 500\u0026ndash;1000 lux on a 16-h light, 8-hr dark cycle, and juvenile \u003cem\u003eHyalella\u003c/em\u003e were fed 3x/week. At the beginning of the test, the overlaying water in the beakers containing control sediment had elevated ammonia levels above recommended levels for Hyalella, so 180 mL of water in each jar was replaced with fresh test water 3x/week. At the end of the test, the final dry weight of \u003cem\u003eHyalella azteca\u003c/em\u003e was noted as well as % survival in each replicate.\u003c/p\u003e \u003cp\u003eImmediately following the toxicity test with \u003cem\u003eHyalella azteca\u003c/em\u003e, a toxicity test with \u003cem\u003eDaphnia magna\u003c/em\u003e was initiated (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) following ecotoxicology protocols Environment and Climate Change Canada (2014). Water (150 ml) from each 1L beaker used in the \u003cem\u003eHyalella azteca\u003c/em\u003e toxicity test was decanted into a 250-mL beaker and 10 \u003cem\u003eDaphnia magna\u003c/em\u003e neonates (ages\u0026thinsp;\u0026lt;\u0026thinsp;24h) were added for a 48-hr exposure test. At the end of the exposure, the number of mobile/surviving daphnids were counted.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec8\" class=\"Section2\"\u003e \u003ch2\u003eAnalysis of Hg, As, Fe and other elements\u003c/h2\u003e \u003cp\u003eBefore adding treatments to sediments and test water, samples of contaminated and artificial control sediment were collected from each beaker as well as the protective capping material. Sediments and protective capping material were dried at a low temperature (40\u0026ndash;60℃), pulverized, and then shipped to Bureau Veritas Minerals Laboratory (Vancouver) for modified aqua regia digestion (1:1:1 HNO3:HCl:H2O) and analysed using ICP-MS for ultra-low determination. Quality controls included sample blanks, duplicates, reagent blanks and Bureau Veritas Certified Reference Materials (OREAS262 and DS11).\u003c/p\u003e \u003cp\u003eTemperature, pH, conductivity, dissolved oxygen, ammonia and alkalinity levels were monitored regularly in the overlaying water of each replicate to ensure acceptable levels for test organisms. At the end of the Hg bioaccumulation test (Day 15), and then again at the end of the \u003cem\u003eHyalella azteca\u003c/em\u003e ecotoxicity tests (Day 34), two 100-ml water samples were obtained from each beaker for total (unfiltered) and dissolved metal(oid) analysis (filtered through 0.45 um Teflon DigiFILTER), and all samples preserved with 1% v/v ultrapure nitric acid. The 100-mL filtered and unfiltered samples were analysed for metal(loids) on ICP-MS by the Analytical Services Unit Laboratory at Queens University. Quality controls included CRMs (EU-H) and in-house reference water samples, all within the 10% range of uncertainty, stock standards, 10% duplicate samples and blanks.\u003c/p\u003e \u003cp\u003eAdditionally, at the end of the \u003cem\u003eHyalella azteca\u003c/em\u003e toxicity tests (day 34), 2 mL porewater was non-destructively sampled from the sediments using syringe-driven Micro-Rhizon\u0026copy; porous samplers. Porewater was collected from three replicates of each treatment, for a total of 27 samples and preserved with 1% v/v ultrapure nitric acid. As the porous part of MicroRhizons samplers have a mean pore size 0.15 \u0026micro;m, additional filtering of the samples before analyzing was unnecessary. Given the small porewater sample volumes, we were only able to analyse Hg via a Direct Mercury Analyzer (DMA) at Saint Mary\u0026rsquo;s University. Results for dissolved Hg analysis were within the 10% range of uncertainty, and each analytical run also included blanks, standard calibration check samples and 10% duplicates.\u003c/p\u003e \u003cp\u003eFiltered and unfiltered water samples, snail and shrimp tissues were analysed for Hg on the Milestone Direct Mercury Analyzer DMA 80.3 at Saint Mary\u0026rsquo;s University. Quality control for each analysis run included blanks, calibration solutions, 10% duplicates and reference materials (National Resources Canada NRCan TORT-3 and DORM-4 for tissue analyses, in-house spiked reference water samples for water).\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec9\" class=\"Section2\"\u003e \u003ch2\u003eData analysis\u003c/h2\u003e \u003cp\u003eData were assessed for normality and equal variance. Since our datasets were non-parametric and varied between 3\u0026ndash;6 replicates, the different treatments were assessed using Kruskal-Wallis (K-W) One Way Analysis of Variance on Ranks with pairwise multiple comparison procedure.\u003c/p\u003e \u003c/div\u003e"},{"header":"Results","content":"\u003cdiv id=\"Sec11\" class=\"Section2\"\u003e \u003ch2\u003enZVI and nZVI coated zeolite characteristics\u003c/h2\u003e \u003cp\u003eSEM scans of the nanoscale zerovalent iron particles in activated NANOFER STAR slurry indicated an approximate grain size of 50\u0026ndash;100 nm, and that particles were highly agglomerated (Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003eb). In contrast, although some agglomeration was still observed, when blended with zeolite during the activation phase, nZVI particles in the NANOFER STAR slurry appeared more evenly spread across the zeolite surface, and appeared with less agglomeration than for nZVI alone (Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003ec). The surface elemental distributions of the three different materials are also illustrated in Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003e. The atomic weight composition for the zeolite- nZVI blend (2c) suggests elemental distributions similar to both zeolite and nZVI (Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003ea and b). However, nZVI particles appear more concentrated in cracks or pores on the zeolite particles surface as illustrated in the different spectras for the zeolite-nZVI blended material in Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003ec.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec12\" class=\"Section2\"\u003e \u003ch2\u003eSediment and porewater characterization\u003c/h2\u003e \u003cp\u003eOf the three different sediments, MP had the highest organic matter content (% LOI) at 9.3% followed by the artificial control sediment at 6.4% and then OS at 3.8%. Clay content was similar between the different sediments ranging between 4.4-5% (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e). The lowest pH was measured in the overlaying water above the OS sediment at 6.4, while the overlaying water pH for artificial control sediment was 7.5. Sulphur content was higher in MP sediment at 4.7% compared with OS (0.03%) and control sediment (0.02%).\u003c/p\u003e \u003cp\u003eTotal As and Hg concentrations in both sets of contaminated sediments (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) exceeded Canadian Council of Ministers for the Environment (CCME) sediment Predicted Effect Level (PEL) guidelines for As (17 mg As/kg) and Hg (0.486 mg Hg/kg). Total sediment As concentration in the untreated MP sediment was very high with an average of 94,211\u0026thinsp;\u0026plusmn;\u0026thinsp;25,437 mg As/kg sediment (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e), while the OS sediments were lower, but still elevated (788\u0026thinsp;\u0026plusmn;\u0026thinsp;560 mg As/kg). This trend is reversed for Hg, with OS sediments (91\u0026thinsp;\u0026plusmn;\u0026thinsp;16 mg Hg/kg) having higher concentrations than the MP sediments (41\u0026thinsp;\u0026plusmn;\u0026thinsp;13 mg/kg). Other metals exceeding CCME PEL in the sediments included Pb, Ni, Fe, Ag and Zn in MP sediment and Pb, Cu, Ag and Zn in OS sediment (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e). The prepared artificial control sediments had metal(oid) concentrations below the CCME PEL guidelines (Table\u0026nbsp;\u003cspan refid=\"Tab1\" class=\"InternalRef\"\u003e1\u003c/span\u003e).\u003c/p\u003e \u003cp\u003ePorewater Hg concentrations (Fig.\u0026nbsp;\u003cspan refid=\"Fig3\" class=\"InternalRef\"\u003e3\u003c/span\u003e) in Old Stamp Mill (OS) sediments treated with \u0026ldquo;B\u0026rdquo;; OSB (0.2\u0026thinsp;\u0026plusmn;\u0026thinsp;0.1\u0026micro;g/L) was significantly lower than for untreated OS (0.9\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2 \u0026micro;g/L) (K-W p\u0026thinsp;=\u0026thinsp;0.004). The porewater Hg concentration from OSB was similar to porewater Hg concentrations in the artificial control sediment. There was no significant difference in Hg porewater concentrations between treated and untreated Muddy Pond (MP) sediment, or between any of the artificial control treatments.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec13\" class=\"Section2\"\u003e \u003ch2\u003eTotal and Dissolved As, Hg and other metal(loids) in overlying water\u003c/h2\u003e \u003cp\u003eTotal and dissolved As, Hg and other metal(loids) in the overlaying water above treated and untreated sediments was measured at day 15 (end of the Hg bioaccumulation test, Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) and then again at day 34 (end of the \u003cem\u003eHyalella azteca\u003c/em\u003e sediment ecotoxicity test, Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e).\u003c/p\u003e \u003cp\u003eIn general, total concentrations of As and Hg in the overlaying water decreased (Figs.\u0026nbsp;\u003cspan refid=\"Fig4\" class=\"InternalRef\"\u003e4\u003c/span\u003e and \u003cspan refid=\"Fig5\" class=\"InternalRef\"\u003e5\u003c/span\u003e) from day 15 to day 34 (Fig.\u0026nbsp;\u003cspan refid=\"Fig1\" class=\"InternalRef\"\u003e1\u003c/span\u003e) as can be expected when particle bound As and Hg settle out in the beakers. On day 34, between 75\u0026ndash;100% of total As in the overlaying water was in the dissolved fraction for all sediments and treatments, with one exception for overlaying water in MPB with 54% of As in dissolved form. Of the total Hg concentrations, between 50\u0026ndash;100% in the overlaying water of all sediments and treatments on day 34 was in dissolved form.\u003c/p\u003e \u003cp\u003eAverage total [As] in the overlaying water above untreated MP sediment was very high on Day 15 (Fig.\u0026nbsp;\u003cspan refid=\"Fig4\" class=\"InternalRef\"\u003e4\u003c/span\u003e) at 1482\u0026thinsp;\u0026plusmn;\u0026thinsp;605 \u0026micro;g As/L and remained relatively high on Day 34 at 1088\u0026thinsp;\u0026plusmn;\u0026thinsp;197 \u0026micro;g As/L. The Day-34 total [As] was significantly higher (K-W, p\u0026thinsp;=\u0026thinsp;\u0026lt;\u0026thinsp;0.001) than what was found in the overlaying water above controls (1\u0026thinsp;\u0026plusmn;\u0026thinsp;0.2 \u0026micro;g As/L), and with the B treatment (3\u0026thinsp;\u0026plusmn;\u0026thinsp;0.8\u0026micro;g/L), but not significantly different than what was found with treatment A (17\u0026thinsp;\u0026plusmn;\u0026thinsp;7\u0026micro;g/L). It is noteworthy that the treatment B drastically decreased total [As] in the overlaying water above MP sediment to below the CCME guideline value for protection of aquatic life at 5 \u0026micro;g/L.\u003c/p\u003e \u003cp\u003eAverage total [As] in the overlaying water above the untreated OS sediment on day 15 was 199\u0026thinsp;\u0026plusmn;\u0026thinsp;267 \u0026micro;g/L (Fig.\u0026nbsp;\u003cspan refid=\"Fig4\" class=\"InternalRef\"\u003e4\u003c/span\u003e). This decreased to 71\u0026thinsp;\u0026plusmn;\u0026thinsp;36 \u0026micro;g/L on day 34. Even so, this was significantly higher (K-W, p\u0026thinsp;=\u0026thinsp;\u0026lt;\u0026thinsp;0.001) than the total concentrations in the overlaying water above control sediment and with the B treatment (2.9\u0026thinsp;\u0026plusmn;\u0026thinsp;2.7\u0026micro;g/L). Treatment B was able to decrease [As] in overlaying water above OSB sediment to concentrations below CCME guideline values for protection of aquatic life. Total [As] concentrations in OS sediment with treatment \u0026ldquo;A\u0026rdquo; measured 9\u0026thinsp;\u0026plusmn;\u0026thinsp;7\u0026micro;g/L. This concentration was not significantly different (K-W, p\u0026thinsp;=\u0026thinsp;\u0026lt;\u0026thinsp;0.001) than in the overlaying water of untreated sediment.\u003c/p\u003e \u003cp\u003eThe highest total [Hg] was found in the overlaying water above the untreated OS sediment on Day 15 (2.4 \u0026micro;g/L\u0026thinsp;\u0026plusmn;\u0026thinsp;1.4). This concentration was significantly higher (K-W, p\u0026thinsp;=\u0026thinsp;0.016) than for those in controls, as well as treatment A (0.87 \u0026micro;g total Hg /L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.96) and B (0.28 \u0026micro;g Hg /L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.13) (Fig.\u0026nbsp;\u003cspan refid=\"Fig5\" class=\"InternalRef\"\u003e5\u003c/span\u003e). On Day 34, the total [Hg] in the overlaying water above the untreated OS sediment had decreased to 0.42 \u0026micro;g Hg/L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.22, indicating the settling out of particulate Hg. This total Hg concentration was no longer significantly different than concentrations found in the overlaying water above controls and treatment A (0.22 \u0026micro;g/L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.04) and B (0.26 \u0026micro;g/L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.09).\u003c/p\u003e \u003cp\u003eThe overlaying water for the untreated Muddy Pond (MP) sediment measured 0.73 \u0026micro;g total Hg /L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.86 on Day 15 and was not significantly higher than overlaying water Hg concentrations in the controls and treatment A and B (Fig.\u0026nbsp;\u003cspan refid=\"Fig5\" class=\"InternalRef\"\u003e5\u003c/span\u003e). On Day 34, the overlaying water Hg concentrations from the untreated MP sediments decreased to 0.2 \u0026micro;g/L\u0026thinsp;\u0026plusmn;\u0026thinsp;0.0.\u003c/p\u003e \u003cp\u003eIt should be noted that total and dissolved Hg in the overlaying water above treated and untreated controls on Day 15 and Day 34 ranged between 0.2\u0026ndash;0.4 \u0026micro;g/L, and this is higher than the CCME guideline value for protection of aquatic life at 0.026\u0026micro;g total Hg/L (Fig.\u0026nbsp;\u003cspan refid=\"Fig5\" class=\"InternalRef\"\u003e5\u003c/span\u003e). On day 15, 100% of the total Hg in the overlaying water in both treated and untreated controls was in the dissolved fraction. The fraction of dissolved Hg remained high (100%) in the untreated control on Day 34, but decreased in treated control sediment to 61% (treatment A), and 52% (treatment B).\u003c/p\u003e \u003cp\u003eTotal overlaying water concentrations of other metal(loid)s which exceeded CCME PEL\u0026rsquo;s included Pb, Cu, and Fe in some cases. In summary, total overlaying water Pb concentrations on Day 15 in untreated control (6\u0026thinsp;\u0026plusmn;\u0026thinsp;6 \u0026micro;g/L), MP (8\u0026thinsp;\u0026plusmn;\u0026thinsp;11 \u0026micro;g/L) and OS (32\u0026thinsp;\u0026plusmn;\u0026thinsp;40\u0026micro;g/L) sediment all exceeded the CCME CWQG Guideline of 1 \u0026micro;g/L. However, on Day 34, total overlaying water Pb concentrations in untreated control (0.5\u0026thinsp;\u0026plusmn;\u0026thinsp;0 g/L) and MP (0.5\u0026thinsp;\u0026plusmn;\u0026thinsp;0 g/L) sediments dropped to below the CCME guidelines, while untreated OS overlaying water concentrations (3\u0026thinsp;\u0026plusmn;\u0026thinsp;2\u0026micro;g/L) still remained above CCME guidelines.\u003c/p\u003e \u003cp\u003eOn Day 15, the overlaying water total [Cu] in untreated controls was 2 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;0.3, while untreated MP was 2 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;0.04, MPA 10 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;9 and MPB 16 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;13. Untreated OS sediment measured 72 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;36, OSA 13 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;10 and OSB 9 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;5. On Day 34, the overlaying water [Cu] for untreated controls, untreated MP and MPA all reported levels of Cu below the CCME CWQG guideline value (2 \u0026micro;g/L). MPB reported a slightly higher concentration at 4 \u0026micro;g\u0026thinsp;\u0026plusmn;\u0026thinsp;2. Overlaying water total [Cu] for untreated OS on day 34 was 28\u0026thinsp;\u0026plusmn;\u0026thinsp;18\u0026micro;g/L, exceeding the CCME guideline value, and was significantly reduced for OSA (by 85%) and OSB (by 81%) treatments.\u003c/p\u003e \u003cp\u003eOverlaying water total [Fe] on Day 34 in untreated MP, MPA and MPB was higher than those on Day 15. Only the MPB (614\u0026thinsp;\u0026plusmn;\u0026thinsp;607 \u0026micro;g/L) overlaying water samples on Day 34 exceeded CCME PEL guideline of 300\u0026micro;g/L, out of which 8% was in the dissolved form.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec14\" class=\"Section2\"\u003e \u003ch2\u003eBioaccumulation of Hg\u003c/h2\u003e \u003cp\u003eAmano shrimp exposed to OS sediments (Fig.\u0026nbsp;\u003cspan refid=\"Fig6\" class=\"InternalRef\"\u003e6\u003c/span\u003e) had lower Hg concentrations in the beakers with treatments A (0.7\u0026thinsp;\u0026plusmn;\u0026thinsp;0.3 mg Hg/kg) and B (0.08\u0026thinsp;\u0026plusmn;\u0026thinsp;0.03 mg Hg/kg) compared to those from untreated OS sediments (6.0\u0026thinsp;\u0026plusmn;\u0026thinsp;5.0 mg Hg/kg), and similar to those found in the control sediments (0.07\u0026ndash;0.08 mg Hg/kg). Amano shrimp exposed to MP sediments (Fig.\u0026nbsp;\u003cspan refid=\"Fig6\" class=\"InternalRef\"\u003e6\u003c/span\u003e) had moderately elevated Hg in untreated MP sediments (2.9\u0026thinsp;\u0026plusmn;\u0026thinsp;3.5 mg Hg/kg) and appeared to have lower Hg concentrations in the beakers with treatment A (0.8\u0026thinsp;\u0026plusmn;\u0026thinsp;0.8 mg Hg/kg) and B (0.2\u0026thinsp;\u0026plusmn;\u0026thinsp;0.3 mg Hg/kg), but those were not statistically significant. The apple snails (\u003cem\u003ePomacea bridgesi\u003c/em\u003e) did not have statistically significant differences in Hg tissue concentrations between any of the contaminated sediment, treated sediment or control sediment, and we observed the snails in the beakers often had their operculum closed during the exposure.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec15\" class=\"Section2\"\u003e \u003ch2\u003eEcotoxicity Sediment Test with Hyalella azteca\u003c/h2\u003e \u003cp\u003e \u003cem\u003eHyalella azteca\u003c/em\u003e 14-day survival in untreated control sediment was satisfactory with an average survival of 94\u0026thinsp;\u0026plusmn;\u0026thinsp;9% (Fig.\u0026nbsp;\u003cspan refid=\"Fig7\" class=\"InternalRef\"\u003e7\u003c/span\u003e), which is in accord with the ECCC Hyalella test method requiring\u0026thinsp;\u0026gt;\u0026thinsp;80% survivability in control treatments (Environment and Climate Change Canada \u003cspan citationid=\"CR10\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). \u003cem\u003eH. azteca\u003c/em\u003e also had high survivability in control sediment with treatment A (96\u0026thinsp;\u0026plusmn;\u0026thinsp;5%) and treatment B (88\u0026thinsp;\u0026plusmn;\u0026thinsp;4%). The average individual dry weight (0.4\u0026thinsp;\u0026plusmn;\u0026thinsp;0.1mg) of control \u003cem\u003eH. azteca\u003c/em\u003e did not change significantly between treated and untreated controls.\u003c/p\u003e \u003cp\u003e \u003cem\u003eH.azteca\u003c/em\u003e 14-d survivability in untreated OS sediment (56\u0026thinsp;\u0026plusmn;\u0026thinsp;22%) was lower than for the control sediments. However, 14-day survivability significantly increased in treatment A (94\u0026thinsp;\u0026plusmn;\u0026thinsp;6%, p\u0026thinsp;=\u0026thinsp;0.006) comparable to survival in control sediments. Survival in treatment B also appeared higher at 90\u0026thinsp;\u0026plusmn;\u0026thinsp;7%. However, this apparent increase in survival was not statistically significant. The average individual dry weight of \u003cem\u003eH. azteca\u003c/em\u003e (0.1\u0026thinsp;\u0026plusmn;\u0026thinsp;0.03 mg) in OS sediments was not significantly different between treated and untreated OS, but was significantly lower than those for control \u003cem\u003eH. azteca\u003c/em\u003e (p\u0026thinsp;=\u0026thinsp;0.008).\u003c/p\u003e \u003cp\u003eThe 14-d survivability of \u003cem\u003eH. azteca\u003c/em\u003e in untreated MP sediments was very low (14\u0026thinsp;\u0026plusmn;\u0026thinsp;9%), with only 7 individual amphipods surviving to the end, as opposed to 45\u0026ndash;50 individuals in all other treatments. There were significant improvements in 14-d survivability for MPA (96\u0026thinsp;\u0026plusmn;\u0026thinsp;5%, p\u0026thinsp;\u0026lt;\u0026thinsp;0.001) and MPB (90\u0026thinsp;\u0026plusmn;\u0026thinsp;10%, p\u0026thinsp;\u0026lt;\u0026thinsp;0.001) treatments. There was no significant difference between the dry weights of \u003cem\u003eH. azteca\u003c/em\u003e exposed to treated and intreated MP sediments, or between MP sediments and controls, but the low survivability of the untreated MP amphipods made it difficult to assess significance.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec16\" class=\"Section2\"\u003e \u003ch2\u003eEcotoxicity Water Test with Daphna magna\u003c/h2\u003e \u003cp\u003eThe 48-h survival of \u003cem\u003eDaphnia magna\u003c/em\u003e in overlaying water taken from above untreated sediments and treated control (98%\u0026plusmn;5) were similar and is acceptable in accordance with the ECCC standard requiring\u0026thinsp;\u0026gt;\u0026thinsp;90% survivability in control treatments (Environment Canada 2014). The 48-h survival (62%\u0026plusmn;34) of \u003cem\u003eD. magna\u003c/em\u003e in water from untreated OS jars varied considerably among replicates (Fig.\u0026nbsp;\u003cspan refid=\"Fig8\" class=\"InternalRef\"\u003e8\u003c/span\u003e). The 48-h survivability improved for OSA (94%\u0026plusmn;9) and OSB (96%\u0026plusmn;6) treatments although the improvement was not statistically significant (p\u0026thinsp;=\u0026thinsp;0.06). The 48-h D. magna survival for water from the untreated MP sediment (84%\u0026plusmn;9) was lower than those for the MPA (98%\u0026plusmn;5) and MPB (96%\u0026plusmn;5) sediment treatments, with MPA being significantly higher (p\u0026thinsp;=\u0026thinsp;0.02).\u003c/p\u003e \u003c/div\u003e"},{"header":"Discussion","content":"\u003cp\u003eThe customized reactive amendment component that was placed on two different contaminated sediments, either below (treatment A), or within (treatment B) the thin protective capping matrix in this study is a novel simple blend of commercially available NANOFER STAR nZVI and fine-grained zeolite (clinoptilolite). When the activated NANOFERSTAR slurry was mechanically blended with zeolite, nZVI particles still agglomerated due to Vander Waals forces and magnetic interactions (He et al., 2007), but the agglomeration of nZVI particles was reduced and they appeared more evenly distributed on the surface of zeolite. This is similar to what was found by Li et al., (\u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e) who used more labour and resource intensive liquid phase reduction and ion exchange procedures to synthesise their zeolite supported nZVI product. Reduction in agglomeration of nZVI and dispersion of nZVI particles of zeolite in their study was more pronounced than in ours. They noted that nZVI particles were homogeneously dispersed on the surface of zeolite, with no obvious agglomeration observed. This more even distribution of nZVI particles on the zeolite is likely due to the liquid phase reduction and ion exchange procedures used to synthesize the zeolite supported nZVI. In our study mechanical mixing and introduction of zeolite during the activation stage of the nZVI was not as effective in dispersing the nZVI on the zeolite. Nevertheless, hydrogen production was much greater in the zeolite/nZVI slurry compared with a control slurry containing nZVI only, indicating higher reactivity in the zeolite/nZVI slurry blend than in the nZVI slurry alone. SEM images also confirm that coating of NANOFERSTAR nZVI on zeolite (Fig.\u0026nbsp;\u003cspan refid=\"Fig2\" class=\"InternalRef\"\u003e2\u003c/span\u003eC) using simple mechanical blending during the activation phase prevented nZVI from agglomerating to some degree, which increased the specific surface area and reactivity of nZVI.\u003c/p\u003e \u003cp\u003e \u003cb\u003eCan Treatment A and B reduce migration of As, Hg, and other metals from contaminated sediments?\u003c/b\u003e \u003c/p\u003e \u003cp\u003eBoth treatment A and B reduced total [As] in the overlaying water above MP and OS sediments but the Treatment B (zeolite-nZVI blend mixed in with a 2cm layer of protective capping) was more successful. Treatment B beakers exhibited decreased total [As] in the overlaying water above MP sediments by 99%, and by 96% in OS sediments, bringing total [As] to below the CCME guideline value for protection of aquatic life (5 \u0026micro;g/L).\u003c/p\u003e \u003cp\u003eNo other studies to the best of our knowledge have attempted to treat As and Hg contaminated wetland sediment with a zeolite-nZVI customized thin reactive capping blend, or zeolite supported nZVI. However, studies have investigated separate components of our blend for direct amendment of contaminated soil and sediment. For example, Li et al. (\u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e) evaluated the bioavailability of metals in aquatic sediments amended with zeolite and found that As porewater concentrations were reduced by 71% after 24 hours. This is in contrast to findings by Kang et al. (\u003cspan citationid=\"CR19\" class=\"CitationRef\"\u003e2016\u003c/span\u003e) who discovered higher As concentrations above sediments amended with zeolite only. Since zeolite has a very low affinity for anionic compounds due to negative surface charges (Jacobs and F\u0026ouml;rstner, 1999), this is not surprising. Nanoscale zero-valent iron (nZVI) on the other hand, has a high anionic adsorption capacity (Mu et al., \u003cspan citationid=\"CR33\" class=\"CitationRef\"\u003e2017\u003c/span\u003e). Gil-D\u0026iacute;az et al., (\u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) found that an nZVI dose of only 5% to soil lead to a 70% decrease of exchangeable As. This is similar to a previous study completed by our laboratory group (Chapman et al. \u003cspan citationid=\"CR7\" class=\"CitationRef\"\u003e2020\u003c/span\u003e), where an 8% dose of nZVI was able to reduce As concentrations in the overlaying water above As-contaminated sediment by 88%. In our current study, in Treatment B where zeolite was combined with nZVI, not only did the total [As] decrease more drastically, but the fraction of dissolved As (potentially more toxic forms) also decreased from 100\u0026ndash;54% above MP sediments. It appears that our zeolite-nZVI blend and a protective capping layer is more successful in reducing As migration from sediments than nZVI or zeolite alone. This is supported by several batch-testing studies assessing zeolite and iron products in As-spiked water. Jeon et al. (\u003cspan citationid=\"CR18\" class=\"CitationRef\"\u003e2009\u003c/span\u003e) investigated the sorption characteristics of arsenic (As(V)) on iron-coated zeolite (ICZ) through batch studies, and they found that As(V) was completely removed within 30 min in a concentration of 2 mg/l, with a 100 g/l dose of ICZ. The adsorption capacity of ICZ for As(V) was 0.68 mg/g. However, arsenite (As(III)) is more prevalent in anoxic or acidic waters common to wetland areas than As(V). As(III) is also more mobile and toxic than arsenate (Sealy, \u003cspan citationid=\"CR39\" class=\"CitationRef\"\u003e2011\u003c/span\u003e). Li et al. (\u003cspan citationid=\"CR26\" class=\"CitationRef\"\u003e2018\u003c/span\u003e) used batch tests to investigate zeolite-supported nanoscale zero-valent iron for adsorption of As(III) in aqueous solution. They found that the maximum adsorption capacity for their zeolite supported nZVI was 11.52 mg As(III)/g, which was much higher than that of zeolite alone. It was suggested that As reacts with zeolite supported nZVI through adsorption mechanisms including electrostatic adsorption, ionic exchange, oxidation, reduction, co-precipitation, and complexation.\u003c/p\u003e \u003cp\u003eOf the two treatments in this study, Treatment B also appeared most effective in preventing Hg release into the overlaying water and porewater from sediments. On day 15, total [Hg] in the overlaying water above Treatment B MP sediment (MPB) had been significantly reduced by 65%, and above \u0026ldquo;B\u0026rdquo; treated OS sediment by 88%. On day 34, [Hg] in the overlaying water in untreated MP and OS beakers were still elevated compared with CCME guideline values but did not differ significantly from total [Hg] found in controls or treatments. This was also the case for dissolved Hg porewater concentrations in treated and untreated MP sediments. However, both Treatments A and B significantly reduced dissolved [Hg] in OS sediment porewater. The dissolved porewater [Hg] was reduced by 35% in OSA and by 73% in OSB. This is similar to findings by Lewis et al. (\u003cspan citationid=\"CR25\" class=\"CitationRef\"\u003e2016\u003c/span\u003e) who investigated nZVI Hg remediation in wetland sediment in a mesocosm set up. They analysed porewater for total Hg and methyl Hg and found that total concentrations decreased in porewater treated with nZVI by approximately 27% and methyl mercury by 42%. They hypothesized that ZVI likely decrease the MeHg concentration via adsorption and not demethylation or inhibition of Hg(II) methylation. However, Gil-D\u0026iacute;az et al. (\u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) found that a dose of 5% of nZVI to contaminated soil did not significantly reduce exchangeable Hg. A higher dose of nZVI (10%) was necessary in their study to achieve reductions of exchangeable-Hg, between 63 and 90% depending on the type of nZVI and soil. Our 73% reduction of dissolved Hg porewater concentrations on addition of treatment \u0026ldquo;B\u0026rdquo; with lower nZVI concentrations than in the Gil-D\u0026iacute;az et al. (\u003cspan citationid=\"CR12\" class=\"CitationRef\"\u003e2017\u003c/span\u003e) study could be due to zeolite\u0026rsquo;s additional adsorption of Hg. Zeolites are proven ion exchange materials where the indigenous (typically sodium) charge balancing cations are readily exchanged with metal cations in solution (Wang et al. \u003cspan citationid=\"CR40\" class=\"CitationRef\"\u003e2010\u003c/span\u003e).\u003c/p\u003e \u003cp\u003eBoth Treatments A and B significantly reduced overlaying water Pb and Cu concentrations above contaminated sediments by at least 77% (A) and 76% (B) for Pb and, 90% (A) and 80% for Cu, to below the CCME guideline values. This is slightly less than for Zhang Xin et al. (\u003cspan citationid=\"CR42\" class=\"CitationRef\"\u003e2010\u003c/span\u003e) who reported 98.8% reduction of Pb(II) in electroplating waste water by using synthesized Kaolin-nZVI..\u003c/p\u003e \u003cp\u003eIn OS sediment beakers, there was no significant difference in total [Fe] between untreated and treated OS sediments, but Treatment B had significantly higher total [Fe] in overlaying water above MPB sediments after 34 days (614\u0026thinsp;\u0026plusmn;\u0026thinsp;607 \u0026micro;g/L), exceeding CCME guideline value (300 ug/L). nZVI can oxidize to Fe\u003csup\u003e2+\u003c/sup\u003e and Fe\u003csup\u003e3+\u003c/sup\u003e rapidly in contact with water and sediment, and very high concentrations of these ions are produced over a short time. This is concerning as even though organisms can tolerate high Fe concentrations, NANOFERSTAR nZVI has been found toxic to \u003cem\u003eDaphnia magna\u003c/em\u003e at concentrations exceeding 0.5 mg/L (Keller et al., 2012). However, the literature is mixed on this, because Yoon et al. (2018) found no acute response of \u003cem\u003eDaphnia magna\u003c/em\u003e when exposed to nZVI with water Fe concentrations\u0026thinsp;\u0026gt;\u0026thinsp;100 mg/L.\u003c/p\u003e \u003cp\u003eIt is noteworthy that our toxicity tests with treated control sediment confirmed that although concentrations of Fe in the overlaying water increased for both Treatments A and B, we did not observe any toxic responses in invertebrates (14 days for \u003cem\u003eHyalella azteca\u003c/em\u003e and 48 hours for \u003cem\u003eDaphnia magna\u003c/em\u003e). The higher total [Fe] in overlaying water above Treatment B in MP sediments (MPB) also was not enough to produce a toxic response in these organisms. This could be because only 8% of the total water [Fe] was dissolved, indicating a lower potential toxicity. It is highly recommended that future assessments of ZVI in-situ treatments for wetland sediments include dissolved as well as total Fe concentrations in overlaying water, as this may clarify some of the confounding results in the literature.\u003c/p\u003e \u003cp\u003eIn general, Treatment B appeared more effective in reducing mobility of metal(oids) from the contaminated sediments to overlaying water compared with Treatment A. It is hypothesized that this could be due to more effective isolation with the 2-cm thick Treatment B capping compared with the 1-cm thick Treatment A capping. In addition, the active ingredients (zeolite/nZVI) had more contact with the overlaying water in Treatment B because the active ingredients were mixed in with the protective capping as opposed to layered underneath as in Treatment A, and which may have contributed to more effective adsorption of Hg and As in the overlaying water.\u003c/p\u003e \u003cdiv id=\"Sec18\" class=\"Section2\"\u003e \u003ch2\u003eCan Treatment A and B reduce bioaccumulation of Hg in invertebrates?\u003c/h2\u003e \u003cp\u003eTotal concentrations of Hg in Amano shrimp exposed to OSB sediment was significantly reduced by 99% from 6 mg Hg/kg in untreated OS sediment to 0.08 mg Hg/kg, which were so low that those were comparable with Hg concentrations in the control shrimp treatments (0.07\u0026ndash;0.08 mg Hg/kg).In the field, LeBlanc et al. (\u003cspan citationid=\"CR24\" class=\"CitationRef\"\u003e2019\u003c/span\u003e) reported that invertebrates (dragonfly larvae, damselfly larvae, and aquatic spiders) collected from the Old Stamp Mill contaminated wetland site (OS) has average dry-weight Hg concentrations of 2 mg/kg, while invertebrates from a nearby uncontaminated reference wetland site had 0.17\u0026ndash;0.24 mg Hg/kg. It appears Treatment B has reduced the bioavailability of sediment Hg to the point that invertebrate Hg concentrations are well below field values even for invertebrates from reference sites. The much lower Hg concentrations in the OSB shrimp is likely due to Hg binding with zeolite-nZVI leading to reduced bioavailability, but also isolation and reduced exposure to the contaminated sediment provided by the thin protective capping. Treatment A for both OSA and MPA as well as MPB also had much lower Hg concentrations in shrimp, but unfortunately this was not statistically significant. Even so, shrimp from OSA, MPA and MPB beakers had similarly low shrimp Hg concentrations (0.2\u0026ndash;0.8 mg Hg/kg) as for OSB shrimp and field data from reference wetlands sites (LeBlanc 2019).\u003c/p\u003e \u003cp\u003eApple snails exposed to treated and untreated wetland sediments did not show a significant difference for Hg concentrations. This is likely due to the fact that this particular snail species (\u003cem\u003ePomacea bridgesi\u003c/em\u003e) was able to avoid exposure to sediments by either climbing along the side of the beakers or closing operculum (trap door) over the shell openings. Due to these avoidance techniques, we conclude that this species of snail is not ideal for assessing bioaccumulation of contaminants from sediments. Lewis et al. (\u003cspan citationid=\"CR25\" class=\"CitationRef\"\u003e2016\u003c/span\u003e) used a different species of snail (\u003cem\u003eLymnaea stagnalis\u003c/em\u003e) (which do not have operculum) in laboratory microcosms and was able to confirm that these snails accumulated less MeHg in sediment treated with ZVI.\u003c/p\u003e \u003c/div\u003e \u003cdiv id=\"Sec19\" class=\"Section2\"\u003e \u003ch2\u003eCan Treatment A and B reduce the toxicity of contaminated sediments?\u003c/h2\u003e \u003cp\u003eThe untreated MP and OS sediments were toxic to \u003cem\u003eHyalella azteca\u003c/em\u003e, with only 14% and 56% surviving the 14-day exposure duration respectively. With As concentrations in MP sediment measuring as high as 94,211 mg/kg and in OS sediment; 788 mg/kg, this was expected and similar to what has been found previously. For example, the 10-day LC50 for \u003cem\u003eHyalella azteca\u003c/em\u003e in As spiked sediment has been previously estimated at 532 mg As/kg (Liber et al. \u003cspan citationid=\"CR29\" class=\"CitationRef\"\u003e2011\u003c/span\u003e). A slightly lower 14-day LC50 was reported by Goulet and Thompson, (2018) at 134 mg As/kg. Interestingly, despite the very high As and Hg sediment concentrations in our study, both Treatments A and B were associated with significantly improved survival rates of \u003cem\u003eHyalella azteca\u003c/em\u003e, close to the survival rates observed in controls. Treatment A was slightly more effective in reducing toxicity of sediments to \u003cem\u003eHyalella azteca\u003c/em\u003e than Treatment B for both contaminated sediment types. It is hypothesized that since Treatment A consisted of a reactive layer of concentrated zeolite/nZVI in direct contact with the contaminated sediment, this may have contributed to more effective binding of labile sediment contaminants to the reactive material reducing As and Hg toxicity. Treatment B contained the same zeolite/nZVI mixture, but it was blended in with the protective capping layer so would not have had the same concentration in contact with contaminated sediments.\u003c/p\u003e \u003cp\u003eOverlying water concentrations of As in untreated beakers was highest in the overlaying water for untreated MP sediments at 1088 \u0026micro;g/L, and water concentrations of Hg was highest in the water overlaying untreated OS sediments at 0.42 \u0026micro;g/L, which corresponds quite well with toxicity findings of Okamoto et al., (2015), who measured an EC50 of \u003cem\u003eDaphnia magna\u003c/em\u003e after 48 hour exposure to be 2400 \u0026micro;g As/L and 0.65 \u0026micro;g Hg/L. Treatments A and B\u0026rdquo; were associated with increased survival rates of \u003cem\u003eDaphnia magna\u003c/em\u003e, with MPA having the highest significance. Survival of \u003cem\u003eDaphnia magna\u003c/em\u003e after 48-hour exposure to the overlaying water of untreated MP and OS sediment on day 34 was 84% and 62% respectively.\u003c/p\u003e \u003cp\u003eTo summarize, our proof-of-concept bench top testing with this reactive zeolite/nZVI amendment blend within and below a thin protective capping shows high potential as an in-situ risk management option for supporting natural recovery of freshwater wetlands impacted by historical gold-mine tailings. The reactive amendments and capping materials are relatively easily processed with commercially-available products. However, to bridge the gap between the laboratory and the real world, it will be necessary to fine-tune component ratios, determine As and Hg adsorption rates of amendment in different sediments, and complete testing using more environmentally realistic mesocosms. It is likely that the nZVI coated zeolite will reduce the thickness of capping needed and improve the overall effectiveness of caps in sequestering contaminants in the sediments, stopping those potentially toxic metal(loids) from migrating into the overlaying aquatic environment. However, concentrations and placement of nZVI in capping materials will have to be carefully assessed before it can be used in the field as the potential environmental impact of nZVI itself is not well known.\u003c/p\u003e \u003c/div\u003e"},{"header":"Declarations","content":"\u003cp\u003e\u003cstrong\u003eAcknowledgements\u0026nbsp;\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eThe authors wish to thank Chrisine Moore with Intrinsik Corp for support with this study, as well as Shane Dalton and Anna Murphy for their assistance with experiments/analysis.\u0026nbsp;\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eFunding details:\u0026nbsp;\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eThis project was supported by a grant from the Nova Scotia Mineral Resources Development Fund (MRDF) IN-58 (2018). The interns who assisted with the project were supported by SMUworks Work-Study funding from Saint Mary\u0026rsquo;s University and a Clean Foundation \u0026ndash; Environment and Climate Change Canada Professional Internship Program. The Scanning Electron Microscopy (SEM) analyses completed for this study were funded by Atlantic Mining Nova Scotia to Dr. Linda Campbell.\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eCompeting interests\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eThe authors have no relevant financial or non-financial interests to disclose.\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eAuthor contributions:\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eBoth authors contributed to the study conception, and experiment was designed by E. Emily V. Chapman. Material preparation, data collection and analysis were performed by E. Emily V. Chapman. The first draft of the manuscript was written by E. Emily V. Chapman and Linda M. Cambell commented on previous versions of the manuscript. Both authors read and approved the final manuscript.\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eEthics approval\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eEthics approval was not required for this study as it did not involve human subjects or vertebrate animals. In Canada, the Canadian Council on Animal Care do not currently require reviews of invertebrate experiments (with the exception of squid and octopus). We followed best national practices for humane invertebrate care and experimental use, including those of Environment Canada and Climate Change federal experimental protocols.\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eConsent to Participate\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eNot applicable as research did not involve human subjects\u003c/p\u003e\n\u003cp\u003e\u003cstrong\u003eConsent to Publish\u003c/strong\u003e\u003c/p\u003e\n\u003cp\u003eNot applicable as research did not involve human subjects\u003c/p\u003e"},{"header":"References","content":"\u003col\u003e\n\u003cli\u003eAlpers, Charles, Michael Hunerlach, Jason May, and Roger Hothem. 2005. \u0026ldquo;Mercury Contamination from Historical Gold Mining in California.\u0026rdquo; 2005. https://pubs.usgs.gov/fs/2005/3014/.\u003c/li\u003e\n\u003cli\u003eAlpers, Charles N., Julie L. Yee, Joshua T. Ackerman, James L. Orlando, Darrel G. Slotton, and Mark C. Marvin-DiPasquale. 2016. \u0026ldquo;Prediction of Fish and Sediment Mercury in Streams Using Landscape Variables and Historical Mining.\u0026rdquo; \u003cem\u003eScience of the Total Environment\u003c/em\u003e 571 (November): 364\u0026ndash;79. https://doi.org/10.1016/j.scitotenv.2016.05.088.\u003c/li\u003e\n\u003cli\u003eAQUAMIN Steering Group. 1996. \u0026ldquo;Assessment of the Aquatic Effects of Mining in Canada: AQUAMIN Supporting Document II: Regional Syntheses.\u0026rdquo; Environment Canada. Https://www.ec.gc.ca/esee-eem/default.asp?lang=En\u0026amp;n=8AF5467D-1.\u003c/li\u003e\n\u003cli\u003eArancibia-Miranda, Nicol\u0026aacute;s, Samuel E. Baltazar, Alejandra Garc\u0026iacute;a, Daniela Mu\u0026ntilde;oz-Lira, Pamela Sep\u0026uacute;lveda, Mar\u0026iacute;a A. Rubio, and Dora Altbir. 2016. \u0026ldquo;Nanoscale Zero Valent Supported by Zeolite and Montmorillonite: Template Effect of the Removal of Lead Ion from an Aqueous Solution.\u0026rdquo; \u003cem\u003eJournal of Hazardous Materials\u003c/em\u003e 301 (January): 371\u0026ndash;80. https://doi.org/10.1016/j.jhazmat.2015.09.007.\u003c/li\u003e\n\u003cli\u003eArshadi, M., M. K. Abdolmaleki, F. Mousavinia, S. Foroughifard, and A. Karimzadeh. 2017. \u0026ldquo;Nano Modification of NZVI with an Aquatic Plant Azolla Filiculoides to Remove Pb(II) and Hg(II) from Water: Aging Time and Mechanism Study.\u0026rdquo; \u003cem\u003eJournal of Colloid and Interface Science\u003c/em\u003e 486 (January): 296\u0026ndash;308. https://doi.org/10.1016/j.jcis.2016.10.002.\u003c/li\u003e\n\u003cli\u003eCerqueira, Beatriz, Emma F. Covelo, Sandra R\u0026uacute;a-D\u0026iacute;az, Purificaci\u0026oacute;n Marcet, Rub\u0026eacute;n Forj\u0026aacute;n, Jos\u0026eacute; Luis R. Gallego, Lukas Trakal, and Luke Beesley. 2022. \u0026ldquo;Contrasting Mobility of Arsenic and Copper in a Mining Soil: A Comparative Column Leaching and Pot Testing Approach.\u0026rdquo; \u003cem\u003eJournal of Environmental Management\u003c/em\u003e 318 (September): 115530. https://doi.org/10.1016/j.jenvman.2022.115530.\u003c/li\u003e\n\u003cli\u003eChapman, Ellen Emily Vanessa, Christine Moore, and Linda M. Campbell. 2020. \u0026ldquo;Evaluation of a Nanoscale Zero-Valent Iron Amendment as a Potential Tool to Reduce Mobility, Toxicity, and Bioaccumulation of Arsenic and Mercury from Wetland Sediments.\u0026rdquo; \u003cem\u003eEnvironmental Science and Pollution Research\u003c/em\u003e 27 (15): 18757\u0026ndash;72. https://doi.org/10.1007/s11356-020-08347-6.\u003c/li\u003e\n\u003cli\u003eDeSisto, Stephanie L., Heather E. Jamieson, and Michael B. Parsons. 2017. \u0026ldquo;Arsenic Mobility in Weathered Gold Mine Tailings under a Low-Organic Soil Cover.\u0026rdquo; \u003cem\u003eEnvironmental Earth Sciences\u003c/em\u003e 76 (22): 773. https://doi.org/10.1007/s12665-017-7041-7.\u003c/li\u003e\n\u003cli\u003eEnvironment and Climate Change Canada,. 2014. \u0026ldquo;Biological Test Method: Acute Lethality of Effluents to Daphnia Magna.\u0026rdquo; Guidance. Aem. May 9, 2014. https://www.canada.ca/en/environment-climate-change/services/wildlife-research-landscape-science/biological-test-method-publications/acute-lethality-effluents-daphnia-magna.html.\u003c/li\u003e\n\u003cli\u003eEnvironment and Climate Change Canada. 2017. \u0026ldquo;Biological Test Method: Test for Survival, Growth and Reproduction in Sediment and Water Using the Freshwater Amphipod Hyalella Azteca.\u0026rdquo; RM/33. Method Development and Applications Unit Science and Technology Branch.\u003c/li\u003e\n\u003cli\u003eGaudet, Lauren. 2022. \u0026ldquo;BIoaccumulation of Contaminants in Amphibians in Historical Gold Mining Areas of Nova Scotia.\u0026rdquo; Honours in Biology, Halifax NS: Saint Mary\u0026rsquo;s University.\u003c/li\u003e\n\u003cli\u003eGil-D\u0026iacute;az, M., J. Alonso, E. Rodr\u0026iacute;guez-Vald\u0026eacute;s, J. R. Gallego, and M. C. Lobo. 2017. \u0026ldquo;Comparing Different Commercial Zero Valent Iron Nanoparticles to Immobilize As and Hg in Brownfield Soil.\u0026rdquo; \u003cem\u003eScience of The Total Environment\u003c/em\u003e 584\u0026ndash;585 (April): 1324\u0026ndash;32. https://doi.org/10.1016/j.scitotenv.2017.02.011.\u003c/li\u003e\n\u003cli\u003eGoulet, Richard R., and Patsy Thompson. 2018. \u0026ldquo;Bioaccumulation and Toxicity of Uranium, Arsenic, and Nickel to Juvenile and Adult Hyalella Azteca in Spiked Sediment Bioassays.\u0026rdquo; \u003cem\u003eEnvironmental Toxicology and Chemistry\u003c/em\u003e 37 (9): 2340\u0026ndash;49. https://doi.org/10.1002/etc.4179.\u003c/li\u003e\n\u003cli\u003eHou, Deyi, Abir Al-Tabbaa, David O\u0026rsquo;Connor, Qing Hu, Yong-Guan Zhu, Liuwei Wang, Niall Kirkwood, et al. 2023. \u0026ldquo;Sustainable Remediation and Redevelopment of Brownfield Sites.\u0026rdquo; \u003cem\u003eNature Reviews Earth \u0026amp; Environment\u003c/em\u003e 4 (4): 271\u0026ndash;86. https://doi.org/10.1038/s43017-023-00404-1.\u003c/li\u003e\n\u003cli\u003eIntrinsik Corp, EcoMetrix, Klohn Crippen Berger and Wood. 2019a. \u0026ldquo;Conceptual Closure Plan For The Historic Goldenville Tailings Areas Goldenville, Nova Scotia Final Report.\u0026rdquo; Consulting. http://www.ap.smu.ca/~lcampbel/NSLands-Historic-Goldenville-Mine-Conceptual-Closure-Plan-2019-07-25.pdf. http://nslands.ca/pdf/projects/NSLands-Historic-Goldenville-Mine-Conceptual-Closure-Plan-2019-07-25.pdf.\u003c/li\u003e\n\u003cli\u003eIntrinsik Corp, EcoMetrix, Klohn Crippen Berger and Wood. 2019b. \u0026ldquo;Conceptual Closure Study For The Historic Montague Mines Tailings Areas. Halifax, Nova Scotia. Final Report.\u0026rdquo; Consulting. NS Lands Inc. http://www.ap.smu.ca/~lcampbel/NSLands-Montague-Mines-Final-Report-2019-07-24.pdf. http://nslands.ca/pdf/projects/NSLands-Montague-Mines-Final-Report-2019-07-24.pdf.\u003c/li\u003e\n\u003cli\u003eJacobs, Patrick H., and Ulrich F\u0026ouml;rstner. 1999. \u0026ldquo;Concept of Subaqueous Capping of Contaminated Sediments with Active Barrier Systems (ABS) Using Natural and Modified Zeolites.\u0026rdquo; \u003cem\u003eWater Research\u003c/em\u003e 33 (9): 2083\u0026ndash;87. https://doi.org/10.1016/S0043-1354(98)00432-1.\u003c/li\u003e\n\u003cli\u003eJeon, Chil-Sung, Kitae Baek, Joon-Kyu Park, Young-Ki Oh, and Sang-Do Lee. 2009. \u0026ldquo;Adsorption Characteristics of As(V) on Iron-Coated Zeolite.\u0026rdquo; \u003cem\u003eJournal of Hazardous Materials\u003c/em\u003e 163 (2): 804\u0026ndash;8. https://doi.org/10.1016/j.jhazmat.2008.07.052.\u003c/li\u003e\n\u003cli\u003eKang, B. Gu, and S. Park. 2016. \u0026ldquo;Evaluation of Sediment Capping Using Zeolite, Montmorillonite, and Steel Slag to Immobilize Heavy Metals.\u0026rdquo; In \u003cem\u003eOCEANS 2016 - Shanghai\u003c/em\u003e, 1\u0026ndash;5. https://doi.org/10.1109/OCEANSAP.2016.7485474.\u003c/li\u003e\n\u003cli\u003eKnox, Anna, Michael Paller, Kenneth Dixon, Danny Reible, and Jesse Roberts. 2011. \u0026ldquo;Final Report Part II - Innovative In-Situ Remediation of Contaminated Sediments for Simultaneous Control of Contamination and Erosion.\u0026rdquo; SERDP Project ER-1501. Department of Defense Strategic Environmental Research and Development Program (SERDP).\u003c/li\u003e\n\u003cli\u003eKong, Xiangke, Zhantao Han, Wei Zhang, Le Song, and Hui Li. 2016. \u0026ldquo;Synthesis of Zeolite-Supported Microscale Zero-Valent Iron for the Removal of Cr(6+) and Cd(2+) from Aqueous Solution.\u0026rdquo; \u003cem\u003eJournal of Environmental Management\u003c/em\u003e 169 (March): 84\u0026ndash;90. https://doi.org/10.1016/j.jenvman.2015.12.022.\u003c/li\u003e\n\u003cli\u003eLeBlanc, Molly. 2019a. \u0026ldquo;Aquatic and Emergent Invertebrate Bioaccumulation and Transfer of Mercury and Arsenic at Historical Gold Mine Tailing Sites of Nova Scotia.\u0026rdquo; Saint Mary\u0026rsquo;s University.\u003c/li\u003e\n\u003cli\u003eLeBlanc, Molly E. 2019b. \u0026ldquo;Bioaccumulation and Transfer of Mercury and Arsenic in Aquatic Invertebrates and Emergent Insects at Historical Gold Mine Tailing Sites of Nova Scotia.\u0026rdquo; MSc, Halifax, Nova Scotia: Saint Mary\u0026rsquo;s University. http://library2.smu.ca/handle/01/28587.\u003c/li\u003e\n\u003cli\u003eLeBlanc, Molly Evelyn, Michael B Parsons, E Emily V. Chapman, and Linda M Campbell. 2019. \u0026ldquo;Review of Ecological Mercury and Arsenic Bioaccumulation within Historical Gold Mining Districts of Nova Scotia.\u0026rdquo; \u003cem\u003eEnvironmental Reviews\u003c/em\u003e, November. https://doi.org/10.1139/er-2019-0042.\u003c/li\u003e\n\u003cli\u003eLewis, Ariel S., Thomas G. Huntington, Mark C. Marvin-DiPasquale, and Aria Amirbahman. 2016. \u0026ldquo;Mercury Remediation in Wetland Sediment Using Zero-Valent Iron and Granular Activated Carbon.\u0026rdquo; \u003cem\u003eEnvironmental Pollution\u003c/em\u003e 212 (May): 366\u0026ndash;73. https://doi.org/10.1016/j.envpol.2015.11.047.\u003c/li\u003e\n\u003cli\u003eLi, Juanying, Chen Zhang, Jianwei Lin, Jie Yin, Jiayan Xu, and Yiqin Chen. 2018. \u0026ldquo;Evaluating the Bioavailability of Heavy Metals in Natural-Zeolite-Amended Aquatic Sediments Using Thin-Film Diffusive Gradients.\u0026rdquo; \u003cem\u003eAquaculture and Fisheries\u003c/em\u003e 3 (3): 122\u0026ndash;28. https://doi.org/10.1016/j.aaf.2018.05.003.\u003c/li\u003e\n\u003cli\u003eLi, Zhangtao, Lu Wang, Jun Meng, Xingmei Liu, Jianming Xu, Fan Wang, and Philip Brookes. 2018. \u0026ldquo;Zeolite-Supported Nanoscale Zero-Valent Iron: New Findings on Simultaneous Adsorption of Cd(II), Pb(II), and As(III) in Aqueous Solution and Soil.\u0026rdquo; \u003cem\u003eJournal of Hazardous Materials\u003c/em\u003e 344 (February): 1\u0026ndash;11. https://doi.org/10.1016/j.jhazmat.2017.09.036.\u003c/li\u003e\n\u003cli\u003eLi, Zhangtao, Lu Wang, Jizi Wu, Yan Xu, Fan Wang, Xianjin Tang, Jianming Xu, Yong Sik Ok, Jun Meng, and Xingmei Liu. 2020. \u0026ldquo;Zeolite-Supported Nanoscale Zero-Valent Iron for Immobilization of Cadmium, Lead, and Arsenic in Farmland Soils: Encapsulation Mechanisms and Indigenous Microbial Responses.\u0026rdquo; \u003cem\u003eEnvironmental Pollution\u003c/em\u003e 260 (May): 114098. https://doi.org/10.1016/j.envpol.2020.114098.\u003c/li\u003e\n\u003cli\u003eLiber, Karsten, Lorne E. Doig, and Suzanne L. White-Sobey. 2011. \u0026ldquo;Toxicity of Uranium, Molybdenum, Nickel, and Arsenic to Hyalella Azteca and Chironomus Dilutus in Water-Only and Spiked-Sediment Toxicity Tests.\u0026rdquo; \u003cem\u003eEcotoxicology and Environmental Safety\u003c/em\u003e 74 (5): 1171\u0026ndash;79. https://doi.org/10.1016/j.ecoenv.2011.02.014.\u003c/li\u003e\n\u003cli\u003eLiu, Tongzhou, Hong Wang, Zhen Zhang, and Da Zhao. 2017. \u0026ldquo;Application of Synthetic Iron-Oxide Coated Zeolite for the Pollution Control of River Sediments.\u0026rdquo; \u003cem\u003eChemosphere\u003c/em\u003e 180 (August): 160\u0026ndash;68. https://doi.org/10.1016/j.chemosphere.2017.04.023.\u003c/li\u003e\n\u003cli\u003eLouise Meunier, Stephen R. Walker, Joanna Wragg, Michael B. Parsons, Iris Koch, Heather E. Jamieson, and Kenneth J. Reimer. 2010. \u0026ldquo;Effects of Soil Composition and Mineralogy on the Bioaccessibility of Arsenic from Tailings and Soil in Gold Mine Districts of Nova Scotia.\u0026rdquo; \u003cem\u003eEnvironmental Science \u0026amp; Technologyi\u003c/em\u003e 44: 2667\u0026ndash;74.\u003c/li\u003e\n\u003cli\u003eMeunier, Louise, Iris Koch, and Kenneth J. Reimer. 2011. \u0026ldquo;Effects of Organic Matter and Ageing on the Bioaccessibility of Arsenic.\u0026rdquo; \u003cem\u003eEnvironmental Pollution\u003c/em\u003e, Nitrogen Deposition, Critical Loads and Biodiversity, 159 (10): 2530\u0026ndash;36. https://doi.org/10.1016/j.envpol.2011.06.018.\u003c/li\u003e\n\u003cli\u003eMu, Yi, Zhihui Air, and Zhang, Lizhi. 2017. \u0026ldquo;Iron Oxide Shell Mediated Environmental Remediation Properties of Nano Zero-Valent Iron.\u0026rdquo; \u003cem\u003eEnvironmental Science: Nano\u003c/em\u003e 4.\u003c/li\u003e\n\u003cli\u003eNANOIRON. 2010. \u0026ldquo;Manual for Preparation of an Aqueous Suspension from Dry Stabilized Iron Powder NANOFER STAR.\u0026rdquo;\u003c/li\u003e\n\u003cli\u003eOkamoto, Akira, Masumi Yamamuro, and Norihisa Tatarazako. 2015. \u0026ldquo;Acute Toxicity of 50 Metals to Daphnia Magna.\u0026rdquo; \u003cem\u003eJournal of Applied Toxicology\u003c/em\u003e 35 (7): 824\u0026ndash;30. https://doi.org/10.1002/jat.3078.\u003c/li\u003e\n\u003cli\u003eOlsta, J. 2007. \u0026ldquo;In-Situ Capping of Contaminated Sediments with Reactive Materials | Ports 2007.\u0026rdquo; In . San Diego, California, S: American Society of Civil Engineers. https://ascelibrary.org/doi/abs/10.1061/40834%28238%2944.\u003c/li\u003e\n\u003cli\u003eParsons, M, K Leblanc, G Hall, A Sangster, J Vaive, and P Pelchat. 2012. \u0026ldquo;Environmental Geochemistry of Tailings, Sediments and Surface Waters Collected from 14 Historical Gold Mining Districts in Nova Scotia.\u0026rdquo; Open file 7150. Geological Survey of Canada.\u003c/li\u003e\n\u003cli\u003eSaunders, Jared R., Loren D. Knopper, Iris Koch, and Kenneth J. Reimer. 2011. \u0026ldquo;Inclusion of Soil Arsenic Bioaccessibility in Ecological Risk Assessment and Comparison with Biological Effects.\u0026rdquo; \u003cem\u003eScience of The Total Environment\u003c/em\u003e 412\u0026ndash;413 (December): 132\u0026ndash;37. https://doi.org/10.1016/j.scitotenv.2011.10.037.\u003c/li\u003e\n\u003cli\u003eSealy, Heather. 2011. \u0026ldquo;Arsenic Mobility and Attenuation in a Natural Wetland at Terra Mine, Northwest Territories, Canada.\u0026rdquo; Degree of Master of Science, Kingston, Ontario, Canada: Queens University.\u003c/li\u003e\n\u003cli\u003eWang, Wei, Minghua Zhou, Qiong Mao, Junjie Yue, and Xu Wang. 2010. \u0026ldquo;Novel NaY Zeolite-Supported Nanoscale Zero-Valent Iron as an Efficient Heterogeneous Fenton Catalyst.\u0026rdquo; \u003cem\u003eCatalysis Communications\u003c/em\u003e 11 (11): 937\u0026ndash;41. https://doi.org/10.1016/j.catcom.2010.04.004.\u003c/li\u003e\n\u003cli\u003eZhang, Chang, Meng-ying Zhu, Guang-ming Zeng, Zhi-gang Yu, Fang Cui, Zhong-zhu Yang, and Liu-qing Shen. 2016. \u0026ldquo;Active Capping Technology: A New Environmental Remediation of Contaminated Sediment.\u0026rdquo; \u003cem\u003eEnvironmental Science and Pollution Research\u003c/em\u003e 23 (5): 4370\u0026ndash;86. https://doi.org/10.1007/s11356-016-6076-8.\u003c/li\u003e\n\u003cli\u003eZhang Xin, Lin Shen, Lu XiaoQiao, and Chen ZuLiang. 2010. \u0026ldquo;Removal of Pb(II) from Water Using Synthesized Kaolin Supported Nanoscale Zero-Valent Iron.\u0026rdquo; \u003cem\u003eChemical Engineering Journal\u003c/em\u003e 163 (3): 243\u0026ndash;48. https://doi.org/10.1016/j.cej.2010.07.056.\u003c/li\u003e\n\u003c/ol\u003e"}],"fulltextSource":"","fullText":"","funders":[],"hasAdminPriorityOnWorkflow":false,"hasManuscriptDocX":true,"hasOptedInToPreprint":true,"hasPassedJournalQc":"","hasAnyPriority":false,"hideJournal":true,"highlight":"","institution":"","isAcceptedByJournal":false,"isAuthorSuppliedPdf":false,"isDeskRejected":"","isHiddenFromSearch":false,"isInQc":false,"isInWorkflow":false,"isPdf":false,"isPdfUpToDate":true,"isWithdrawnOrRetracted":false,"journal":{"display":true,"email":"[email protected]","identity":"researchsquare","isNatureJournal":false,"hasQc":true,"allowDirectSubmit":true,"externalIdentity":"","sideBox":"","snPcode":"","submissionUrl":"/submission","title":"Research Square","twitterHandle":"researchsquare","acdcEnabled":true,"dfaEnabled":false,"editorialSystem":"","reportingPortfolio":"","inReviewEnabled":false,"inReviewRevisionsEnabled":true},"keywords":"Toxicity, mercury, arsenic, sediment, in-situ risk management","lastPublishedDoi":"10.21203/rs.3.rs-3894488/v1","lastPublishedDoiUrl":"https://doi.org/10.21203/rs.3.rs-3894488/v1","license":{"name":"CC BY 4.0","url":"https://creativecommons.org/licenses/by/4.0/"},"manuscriptAbstract":"\u003cp\u003eLegacy gold mine tailings from the 1800\u0026rsquo;s in Nova Scotia, Canada have elevated mercury (Hg) and arsenic (As) concentrations. Tailings, were slurried into wetlands without treatment. Over a century later, those impacted wetlands are still at risk and innovative in-situ treatment approaches to support natural biological and chemical recovery are needed. Here we report results of our proof-of-concept laboratory study with a customized reactive thin layer capping to limit mobility, bioaccumulation and toxicity of Hg and As in wetland sediment impacted by legacy tailings. The customized reactive amendment is a blend of NANOFER STAR nanoscale zero valent iron (nZVI) and fine-grained zeolite (clinoptilolite) inserted either below, or within a thin cap (silica sand, bentonite and zeolite) and placed over contaminated wetland sediments in beakers. Due to the high concentrations of Hg and As in sediments, invertebrates (\u003cem\u003eHyalella azteca\u003c/em\u003e, \u003cem\u003eDaphnia magna\u003c/em\u003e and \u003cem\u003eCaridina multidente)\u003c/em\u003e exposed to untreated wetland sediment exhibited high mortality and bioaccumulation of Hg. The reactive capping applications improved the survival of \u003cem\u003eH. azteca\u003c/em\u003e and \u003cem\u003eD. magna\u003c/em\u003e similar to the survival rates seen in our clean control sediment. Bioaccumulation of Hg was also reduced in \u003cem\u003eC. multidente\u003c/em\u003e exposed to the treated sediment compared to the untreated sediment. Furthermore, total [Hg] and [As] in the overlaying water of treated contaminated sediments were reduced by 88% and 99% respectively. Our proof-of-concept testing of this reactive capping blend shows potential for managing and supporting natural recovery of wetlands impacted by historical gold-mine tailings.\u003c/p\u003e","manuscriptTitle":"Evaluation of a customized reactive nanoscale-zero-valent iron and zeolite thin capping blend for enhancing natural recovery of wetlands impacted by contaminated legacy gold mine tailings","msid":"","msnumber":"","nonDraftVersions":[{"code":1,"date":"2024-03-25 10:49:31","doi":"10.21203/rs.3.rs-3894488/v1","editorialEvents":[{"type":"communityComments","content":0}],"status":"published","journal":{"display":true,"email":"[email protected]","identity":"researchsquare","isNatureJournal":false,"hasQc":true,"allowDirectSubmit":true,"externalIdentity":"","sideBox":"","snPcode":"","submissionUrl":"/submission","title":"Research Square","twitterHandle":"researchsquare","acdcEnabled":true,"dfaEnabled":false,"editorialSystem":"","reportingPortfolio":"","inReviewEnabled":false,"inReviewRevisionsEnabled":true}}],"origin":"","ownerIdentity":"fe31ccbb-ab12-4552-a360-aa78d92c3ede","owner":[],"postedDate":"March 25th, 2024","published":true,"recentEditorialEvents":[],"rejectedJournal":[],"revision":"","amendment":"","status":"posted","subjectAreas":[],"tags":[],"updatedAt":"2024-04-15T21:16:33+00:00","versionOfRecord":[],"versionCreatedAt":"2024-03-25 10:49:31","video":"","vorDoi":"","vorDoiUrl":"","workflowStages":[]},"version":"v1","identity":"rs-3894488","journalConfig":"researchsquare"},"__N_SSP":true},"page":"/article/[identity]/[[...version]]","query":{"redirect":"/article/rs-3894488","identity":"rs-3894488","version":["v1"]},"buildId":"qtupq5eGEP_6zYnWcrvyt","isFallback":false,"isExperimentalCompile":false,"dynamicIds":[84888],"gssp":true,"scriptLoader":[]}