British Columbia Mine Reclamation Symposium

Investigations of metal mobility and bioavailability in spill-affected lakes Stecko, Pierre; Batchelar, Katharina; Hughes, Colleen; Anglin, Lyn 2018

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INVESTIGATIONS OF METAL MOBILITY AND BIOAVAILABILITY IN SPILL-AFFECTED LAKES  Pierre Stecko, M.Sc, EP, R.P.Bio.1 Katharina Batchelar, M.Sc.1 Colleen Hughes, EP2 ‘Lyn Anglin, Ph.D., P.Geo.3  1 Minnow Environmental Inc. 101 - 1025 Hillside Ave. Victoria, B.C.  V8T 2A2  2 Mount Polley Mining Corporation PO Box 12 Likely, B.C.  V0L 1N0   3 Imperial Metals Corporation 200 – 580 Hornby St. Vancouver, B.C.  V6C 3B6  ABSTRACT  A foundational failure of the perimeter embankment of the Mount Polley Tailings Storage Facility (TSF) on August 4th 2014 resulted in a breach that released approximately 25 million cubic meters of debris (water and solids that consisted of tailings, construction materials, and scoured sediment and soil).  The debris flowed into Polley Lake, along the length of Hazeltine Creek, and into Quesnel Lake.  Sediment quality investigations indicated that sediments within the debris path contained concentrations of copper that were consistently greater than baseline and/or reference concentrations and British Columbia’s upper working sediment quality guideline, concentrations of arsenic and iron that were greater than baseline and/or reference concentrations and British Columbia’s lower working sediment quality guidelines, and concentrations of manganese that were greater than reference concentrations (in affected areas of Quesnel Lake but not Polley Lake) and British Columbia’s lower working sediment quality guideline.  All of the elevated metals have been previously observed at concentrations greater than working sediment quality guidelines at un-impacted areas of the Polley Lake and Quesnel Lake watersheds.  A number of lines of investigation were carried out to characterize the mobility and potential bioavailability of metals within debris-impacted lake sediments, including selective chemical extractions, characterization of metal binding constituents (e.g., organic matter and volatile sulphides), application of diffusive gradients in thin films (DGT) devices, sediment toxicity testing, and benthic invertebrate community monitoring.  Selective chemical extractions indicated that concentrations of the metals most elevated in the debris path were predominantly present in non-mobile forms and that concentrations in forms that are considered to be potentially mobile and/or bioavailable (exchangeable, carbonate, and easily reducible forms) were generally below lower working sediment quality guidelines.  For copper, this finding was aligned with its known presence in residual alumino-silicate minerals as well as with years of geochemical test work conducted on Mount Polley tailings.  DGT devices, which measure free and labile metals, indicated low free and labile concentrations that have been stable and, for copper, were much lower than those associated with effects observed elsewhere.  Standard sediment toxicity tests of Chironomus dilutus and Hyalella azteca documented no unacceptable responses to survival or growth of either organism in Polley Lake.  Sediment toxicity test results for Quesnel Lake were complicated by the very low organic carbon content of debris-impacted sediments and thus observed survival and growth responses appeared to be at least partly associated with adverse physical conditions.  Benthic invertebrate community monitoring has documented recovery in Polley Lake and in the breach-influenced area of Quesnel Lake, as apparent in endpoints of density (organisms per square meter), taxon richness (the number of different organisms present), and taxon composition.  In Polley Lake, recovery of these endpoints was to conditions similar to pre-breach and reference.  Recoveries were supported by sediment transplant experiments in deep depositional areas of Quesnel Lake (cross-over design) which indicated that debris-impacted sediments could be readily colonized when moved to reference areas.  Although the lines of evidence were not completely aligned, overall integration of the weight of evidence collected from 2014 to 2017 suggests limited mobility and bioavailability of metals in debris-impacted sediments of Polley and Quesnel lakes.   KEY WORDS  Sediment quality, metal mobility, metal bioavailability, selective chemical extraction, diffusive gradients in thin films, sediment toxicity testing, benthic invertebrate community monitoring   INTRODUCTION  Failure of a glacio-lacustrine unit underlying the perimeter embankment of the Mount Polley Tailings Storage Facility (TSF) resulted in a breach on August 4th 2014.  The TSF breach released approximately 25 million cubic meters of debris (water and solids that consisted of tailings, construction materials, and scoured sediment and soil), which flowed into Polley Lake, along the length of Hazeltine Creek, and into Quesnel Lake (Figure 1).  The Mount Polley Mining Corporation (MPMC) immediately initiated stabilization work and impact assessment, which included terrestrial and aquatic environments.  Sediment quality assessment is one component of the integrated aquatic environmental impact assessment.  This ongoing assessment has included characterization of sediment physical condition, sediment chemical condition, a number of investigations to assess the mobility and bioavailability of metals in breach-impacted sediment, as well as sediment toxicity testing and benthic invertebrate community assessment.  This paper is focussed on sediment quality of breach-affected lakes and on the investigations of metal mobility and bioavailability.  The fundamental objectives of this work were to characterize the physical and chemical condition of sediment impacted by the TSF breach, and to advance understanding of sediment metal stability or mobility and potential bioavailability in the impacted lake areas.METHODS Sampling Sediments were sampled at five replicate stations per area in Polley Lake, Quesnel Lake, and reference areas (Figure 2), using a either a core sampler (a Kajak-Brinkhurst corer or a Tech Ops corer) or a ponar grab sampler in accordance with technical guidance for sediment sampling outlined in the British Columbia (BC) Field Sampling Manual (BCMOE 2013) and the Metal Mining Technical Guidance for Environmental Effects Monitoring (EEM; Environment Canada 2012).  Every sample was a composite at least 3 sub-samples of the top 3 centimeters (cm) of sediment.  The top 3 cm of sediment from each core was extruded into an extrusion collar marked at 1 cm intervals and cut at the 3 cm mark using a core slicer.  The top 3 cm of sediment from each accepted grab was collected by first placing the sediment in a tote and, if the sample was deemed acceptable (i.e., the sampler was full to each edge and representative of surface sediment), surficial material to a depth of 3 cm was transferred to a second tote using a stainless steel spoon.  Following the collection of three acceptable cores or grabs, the three sub-samples were homogenized and transferred into 250 mL glass jars labeled with the project number, sample location, and collection date.  Duplicate (split) sediment samples were collected at a frequency of 10% of the total number of samples for quality assurance/quality control (QA/QC) purposes.  Supporting information collected at each sediment monitoring station included GPS (Geographic Positioning System) coordinates, sampling depth, Secchi depth, field meter measurements of temperature, specific conductance, dissolved oxygen and pH, photographs of sediment samples, notes of the presence or absence of aquatic vegetation, and other physical observations (sediment texture, colour, density, etc.).  Sampling equipment was rinsed between stations using site water.  Immediately after collection, sediment samples were placed into a cooler with ice packs, where they were maintained cool prior to transport to the field laboratory where they were transferred to a refrigerator and held until shipment to the analytical laboratory.             Additional sediment samples for selective extraction analysis (SEA), for the determination of acid volatile sulphides (AVS) and simultaneously extracted metals (SEM), and for sediment toxicity testing were sampled as outlined above, with one modification for the SEM/AVS samples.  Due to the sensitivity of the sulphide ion measured in AVS analyses to oxidation (USEPA 1991), the SEM/AVS samples were placed in jars without homogenization, the jars were capped between sub-samples, and the jars were filled completely to ensure no headspace.  Furthermore, at the field laboratory, the SEM/AVS jars were placed in nitrogen purged bags to reduce the possibility of oxidation.    Diffusive Gradients in Thin Films devices (DGT devices) are small devices that contain a membrane filter, a gel diffusion layer and a gel binding layer (e.g., Davison and Zhang 1994).  The devices are designed to accumulate labile (free and weakly complexed) substances by diffusion in a controlled manner.  DGT devices used in this evaluation were optimized for metals and include a 0.45 µm Millipore™ cellulose nitrate membrane, a polyacrylamide gel diffusion layer, and a mixed binding layer composed of Chelex-100™ and Metsorb™ binding resins, all housed in a small polypropylene container approximately two inches in diameter (Panther et al. 2014).  Only the membrane is exposed to the sampling environment (as a “window” in the DGT device).  Labile metals diffuse through the membrane and gel diffusion layer and are captured in the resin.  Following deployment for a time period that results in accumulation to detectable concentrations (but not to resin saturation, which can be calculated based on water chemistry data and avoided), the DGT devices are retrieved and the resin removed, digested, and analyzed.  Analytical results are expressed as the mass of metal accumulated in the resin, which can also be used to back-calculate average aqueous labile metal concentrations during deployment.  DGT devices were installed in the lower water column of Polley Lake, Quesnel Lake, and reference areas.  Five replicate DGT devices were placed in exposed areas and three replicates were placed in reference areas.  The DGT devices were installed in the lower water column approximately one meter above the sediment surface and left in place for a total period of 35 to 50 days.  Upon retrieval, each DGT device was gently rinsed with metal free water and placed into a labelled glass soil jar with a few drops of metal free water to prevent desiccation.  Jars were then placed into clean Ziploc™ bags and then into a dedicated cooler with ice packs, where they were maintained cool prior to transport.  Supporting information collected at each DGT deployment location included all those listed for sediment sampling, and the additional collection of supporting water samples for total and dissolved metals, the latter prepared by field filtration using a 0.45 µm syringe filter.  Water samples were placed into a dedicated cooler with frozen ice packs, and kept cool prior to transport to the field laboratory.   Benthic invertebrate sampling for community analysis was completed in accordance with provincial and federal technical guidance (i.e., BCMOE 2013; Environment Canada 2012).  Samples were collected using a ponar grab sampler, with only high quality, comparable samples retained.  Acceptable samples were rinsed from the ponar into a tub to ensure the complete removal of all material   The sample material was then placed into a 250 µm mesh sieve bag and sieved free of all material less than 250 µm in diameter.  At each station, a composite sample of five acceptable grabs was collected (0.116 m2 of bottom area in total) to ensure that each sample was representative of the station.  As for sediment, five replicate samples were collected at each area Figure 2).  After sieving, the retained material (sample) from the five composited grabs was carefully transferred to 1 L wide mouth plastic jars using a wash bottle while working over a plastic tub to avoid any potential loss of organisms.  Each jar received an internal and external label with the station number, area identifier, project number, date, and the initials of the field personnel.  Each sample was preserved with buffered formalin solution to a concentration of 10% within six hours of collection.    Analytical Methods Sediment samples were shipped to the analytical laboratory (ALS Environmental, Burnaby BC) at a minimum frequency of bi-weekly.  Prior to shipment, samples were placed in a cooler with frozen ice packs and a chain-of custody form was prepared and packed with the samples.  Coolers were shipped overnight for next day delivery.  Chemical analyses were completed using methods consistent with those specified in the British Columbia Environmental Laboratory Manual (BCMOE 2015) and included moisture content, pH, particle size distribution, total organic carbon (TOC), total nitrogen, total sulphur, and total metal concentrations.  All analyses except for particle size distribution were completed on the silt/clay fraction (<63 µm diameter) in accordance with recent recommendations by BCMOE (2016).  Analysis of metals was based on the BC Strong Acid Leachable Metals digestion (SALM; BCMOE 2016).  Analyses of pH were completed by electrode, particle size analysis by sieve and pipette, total organic carbon by combustion and titration (total carbon minus total inorganic carbon), total nitrogen and total sulphur by combustion, metals by collision reaction cell inductively coupled plasma – mass spectrometry (USEPA Method 6020A; USEPA 1998), and mercury by cold vapor atomic fluorescence spectrophotometry or atomic absorption spectrophotometry (USEPA Method 245.7; USEPA 2005).  SEA was completed according to a procedure modified from the sequential extraction scheme of Tessier et al (1979).  The “Tessier Procedure” involves five sequential extraction steps employing extractants of increasing strength: 1) magnesium chloride (exchangeable and adsorbed metals); 2) sodium acetate (carbonate metals); 3) hydroxylamine hydrochloride in acetic acid (easily reducible and iron oxide associated metals); 4) nitric acid then ammonium acetate (organic bound metals); and 5) hydrochloric acid and nitric acid (residual metals).  Metals released in each digest were then determined by collision cell ICP-MS (USEPA Method 6020A; USEPA 1998).  The analysis of acid volatile sulphides (AVS) and simultaneously extracted metals (SEM) was performed according to the USEPA Method EPA-821-R-91-100 (USEPA 1991).  This analysis consisted of the cold extraction of sediment in hydrochloric acid (6M HCl at room temperature), reported to release readily extracted reactive sulphides (AVS).  The concentrations of the reactive metals that are concurrently extracted during the cold acid extraction were also measured (SEM).    DGT devices were shipped in coolers with ice packs to the analytical laboratory (Maxxam Analytics, Mississauga ON) within two days of collection.  Laboratory analysis of the DGT devices involved dismantling the devices followed by digestion of the resin to determine total metal concentrations by ICP-MS (USEPA Method 6020A; USEPA 1998).  Aqueous DGT-labile metal concentrations were subsequently calculated by the analytical laboratory based on total metal in the resin, mean water temperature during deployment at each area, total deployment time, and manufacturer-supplied peer-reviewed diffusion coefficients, which, in turn, are based on Fick’s first law of diffusion (Zhang and Davison 1995; Panther et al. 2014).    Toxicity testing was completed at a CALA accredited toxicity testing laboratory (Nautilus Environmental, Burnaby BC).  Testing was completed using a standard 10 day test of the survival and growth of Chironomus dilutus (a midge larva; Environment Canada 1997) and a standard 14 day test of the survival and growth of Hyalella azteca (an  amphipod; Environment Canada 2013).  In accordance with recent recommendations for toxicity testing, test replicates (n = 5) were field-collected replicates.  In addition to the exposed area field replicates and the reference area field replicates, a laboratory control (beach sand in accordance with the standard biological test methods) was also included (also set up with five laboratory replicates).  Lastly, preserved benthic invertebrate community samples were stored at room temperature until shipment to the taxonomy laboratory (Cordillera Consulting, Summerland BC) for taxonomic identification and enumeration.  Taxonomic identification of benthic invertebrate community samples was completed to the lowest practical level (LPL) by a qualified laboratory using standard methods that incorporate QA/QC measures (e.g., Environment Canada 2012).   Data Interpretation Upon receipt of data from the analytical laboratories, data were summarized by calculating mean, median, standard deviation, standard error, minimum, maximum, 5th percentile, and 95th percentile for each analyte.  Sediment quality data were evaluated in comparison to baseline and/or reference concentrations and British Columbia Working Sediment Quality Guidelines for the protection of aquatic life (WSQGs; BCMOE 2018).  Analytes with concentrations greater than baseline/reference concentrations and guidelines were the focus of more detailed interpretation, including graphical comparisons to reference and historical data, and formal statistical contrasts of exposed versus baseline/reference concentrations and/or over time by analysis of variance (ANOVA).  For these analyses, all assumptions were tested, and, if necessary, data transformed to satisfy assumptions of normality and homogeneity of variance.  In instances where the residuals of the ANOVA model did not meet the assumptions of normality and homogeneity of variances, a non-parametric Kruskal-Wallis test was conducted.  When the overall results of ANOVA or the Kruskal-Wallis test were significant, differences between specific areas were evaluated using post-hoc contrasts for ANOVA or Dunn’s test for the Kruskal-Wallis test.  A Bonferroni correction on the p-values was used to conserve the overall Type I error rate.  For some comparisons, data were available to support before-after control-impact comparisons using Two-Way ANOVA (with factors Area and Year), with data treatment as indicated above, and if assumptions were not met, the two-way analysis was conducted on ranks (equivalent to a non-parametric test).     Total concentrations of the simultaneously extracted metals (SEM) were compared to those of acid volatile sulphides (AVS) on a molar basis to determine whether SEM were in excess of AVS.  This was achieved by calculating both the ratio of total SEM/AVS, and the difference between total SEM and AVS [(Ʃ total SEM)-(AVS)].  If the former is less than one (unity) or the latter is less than zero, AVS is in excess of SEM and an absence of these metals in the potentially free (bioavailable) form in interstitial water (porewater) is implied, and an absence of toxicity is predicted (e.g., McGrath et al. 2002).    Sediment toxicity test results (survival and growth of C. dilutus and H. azteca) were interpreted based on a combination of effect size and statistical contrasts against the matched reference area and against the laboratory control.  Results were then evaluated using a stepwise procedure modified from Bay et al. (2014).  This procedure results in categories of: 1) no unacceptable response - a mean response within 20% of the mean field reference response (i.e., survival or growth were ≥80% of the field reference); 2) low to moderate response - a mean response of between 20% and 50% of the mean field reference response (i.e., survival or growth were 50% to 80% of the field reference); and 3) a large response - a mean response of >50% of the mean field reference response (i.e., survival or growth were <50% of the field reference).  Low and moderate responses were distinguished based on whether the results differed significantly from reference (moderate response) or not (low response).  Application of a 20% reduction in survival relative to the field reference to define acceptability is consistent with thresholds used elsewhere (e.g., BCMELP 1998; Government of Canada 2010; Bay et al. 2014).    Benthic invertebrate communities were evaluated using metrics of organism density (organisms per m²), taxonomic richness, Simpson’s Index of Diversity, and Simpson’s Index of Evenness (calculated as in Smith and Wilson 1996; Environment Canada 2012).  These indices were calculated at the lowest practical level (LPL) of taxonomy for the 250 µm fraction of all samples.  Densities and relative densities of the most abundant taxa were also calculated and compared, and multivariate statistical techniques (ordination methods) were used to assist in the identification of differences in community composition among areas (i.e., which stations or areas had distinct benthic communities and how these benthic communities differ among stations).  All metrics were evaluated by formal statistical contrasts of exposed versus baseline/reference concentrations and/or over time by analysis of variance (ANOVA) as described for sediment quality.  In cases where exposed versus reference or temporal differences in benthic invertebrate community results were noted, potential relationships between benthic invertebrate community metrics and physical and chemical conditions (e.g., sediment quality) were explored using correlation analysis.  Evidence of causation obtained from correlation analysis was carefully considered in light of the strength, range, and continuity of the relationships as well as mechanistic knowledge/understanding of potential cause.  RESULTS AND DISCUSSION Conventional sediment quality (digestion with strong acid in accordance with standard BC protocols), indicated that copper was elevated relative to baseline and/or reference concentrations and to concentrations greater than the upper WSQG in debris-influenced areas of Polley Lake and Quesnel Lake, and that arsenic and iron were consistently elevated relative to baseline and/or reference concentrations and to concentrations greater than the lower WSQG in these areas.  Manganese was consistently elevated relative to reference concentrations in Quesnel Lake only (not Polley Lake) where it was present and was consistently greater than the lower WSQG.  Concentrations of nickel and zinc were elevated relative to reference and lower WSQGs at some locations only (Tables 1 and 2).  Concentrations of copper, manganese, arsenic, and iron were naturally elevated in Polley Lake prior to the breach and in the reference lake (Bootjack Lake).  These natural elevations are consistent with elevations that have been noted in regional geochemical databases (particularly for arsenic, which has been observed at concentrations greater than the upper WSQG in numerous un-impacted creeks in the Quesnel Lake watershed).  The TSF breach resulted in increases of approximately 2-fold for copper (Figure 3) and arsenic, and 1.1-fold for iron and manganese.  Concentrations of copper in debris-influenced areas of Quesnel Lake were substantially elevated relative to reference areas (approximately 18-fold; Figure 3), whereas arsenic, iron, and manganese were influenced to lower magnitudes (2-fold, 1.3-fold, and 2-fold, respectively).  Copper concentrations substantially greater than upper WSQGs would be conventionally considered “likely to cause severe effects” (BCMOE 2018).  However, the combination of natural concentrations in excess of upper WSQGs and the fundamentally different characteristics of the debris-impacted sediments (i.e., tailings and debris as opposed to typical lake sediment) require additional techniques to evaluate potential effects (as presented below).  The TSF breach also had a notable influence on the concentrations of total organic carbon (TOC) in sediment (as the debris was largely inorganic), resulting in approximately half of pre-breach/reference TOC in Polley Lake and approximately one-tenth of reference in the most impacted area of Quesnel Lake (Figure 4).  Since 2014, concentrations of TOC have increased slightly in the debris-influenced deep area of Quesnel Lake (from 0.17% in 2014 to 0.35% in 2016; Figure 4).  The TOC concentrations observed in the impacted deep area of Quesnel Lake were in a range that has been shown to adversely affect some toxicity test organisms.  Specifically, Suedel and Rodgers (1994) reported Chironomus intolerance to sediment with organic content less than 1% and poor survival when organic content was 0.5% or lower, even in tests where food was added.    Conventional sediment quality investigations typically employ strong acid extraction to solubilize metals for analytical determination (reported as Strong Acid Leachable Metals - SALM).  However, digestion with strong acid is aggressive and does not reflect the fraction of sediment-associated metal that is mobile and bioavailable in natural water bodies.  Additional information was provided by selective extraction analysis,      which indicated that the majority of copper was in “residual and/or organic” phases (i.e., it can only be extracted using a strong acid [aqua regia] or using a slightly less aggressive acid and a strong oxidizer [nitric  acid and ammonium acetate]; Figure 5).  This finding is consistent with mineralogy and geochemical investigations undertaken by the mine, which have shown that a significant quantity of metal associated with the debris is non-sulphide (alumino-silicate), is not acid generating, and that leaching under neutral to alkaline conditions is low.  This is further supported by the strong relationship between copper and aluminum in impacted sediment, but which is absent in reference sediments (Figure 6).             Concentrations of several key sediment metal binding constituents differed substantially between the impacted lake areas (i.e., in Polley Lake versus the impacted areas of Quesnel Lake).  In Polley Lake, organic carbon content decreased moderately (was roughly halved, to an average of 7.4%; Table 1; Figure 4) and sulphide concentrations were in “normal” ranges.  In impacted areas of Quesnel Lake, organic content and sulphide concentrations were both very low (less than 0.35% and less than detection of 0.2 µmol/g, respectively).  Application of the SEM/AVS methodology indicated that AVS was in excess of SEM in Polley Lake sediments, suggesting that the acid volatile metals are predominantly bound by the acid volatile sulphides (Figure 7).  In Quesnel Lake, AVS concentrations were too low to support this; metals are not predominantly bound to AVS in Quesnel Lake and could be mobile or bioavailable if other binding materials are not present (however, high iron and manganese concentrations in an oxic environment suggest that oxides of these metals are also present to bind free metals).   The application of DGT devices indicated that DGT-labile concentrations of copper (free and weakly complexed copper) were a small fraction of total and dissolved concentrations (9% to 12% and 16% to 18%, respectively; Figure 8).  Although the fraction of DGT-labile copper was slightly greater at impacted areas than at reference areas (approximately 7% of dissolved copper), DGT-labile copper concentrations (maximum 0.58 µg/L) were much lower than concentrations previously shown to cause toxicity to a copper-sensitive test organism (the cladoceran Ceriodaphnia dubia; 15 to 24 µg/L; Martin and Goldblatt 2007).  The application of DGT devices has also indicated temporal stability in DGT-labile copper concentrations from 2015 to 2017 (Figure 9).   Sediment toxicity testing indicated no unacceptable responses (>20%) to survival or growth of the two test organisms (Chironomus dilutus and Hyalella azteca) in standard exposures (10-day and 14-day tests, respectively) in Polley Lake (Table 3).  In Quesnel Lake, moderate to high responses were observed in toxicity tests (Table 3).  Despite some chemical conditions in the toxicity testing vessels that were highly dissimilar from those observed in-situ (e.g., concentrations of overlying copper that were 40-times higher than in-situ [35 µg/L in test vessels versus 0.87 µg/L in-situ], as has been observed elsewhere; e.g., Fetters et al. 2016), multivariate statistical evaluation of the toxicity test results suggested that the observed responses were best described by a combination of low TOC and elevated copper.  It is noted that the identification of TOC and copper as best descriptors of responses is not surprising given that these were the variables most influenced by the breach.  The relationships do not necessarily indicate cause.  However, the observed responses to Chironomus did correspond well to a threshold TOC level below which organisms do not survive or grow well (Suedel and Rodgers 1994).  Benthic invertebrate community monitoring in 2014 indicated significant impact to benthic invertebrate communities of Polley Lake and Quesnel Lake in association with the debris inflows.  In Polley Lake, the recovery of the benthic invertebrate community has been rapid.  In 2016, benthic invertebrate density and      taxon richness were similar to pre-breach and reference (Figure 10), as were most benthic invertebrate community metrics.  In the impacted areas of Quesnel Lake, some improvement has been documented, but both density and taxon richness remained lower than reference (Figure 11), and community composition differed from reference on the basis of more Chirononidae (midge larvae) and fewer oligochaetes (aquatic worms).  This difference is consistent with their dispersal capacities (Chironomidae are stronger dispersers than oligochaetes).  The recovery observed in Quesnel Lake was supported by sediment transplant experiments (2-way controlled design) which showed that impacted sediment can be recolonized when transferred to a reference area (which provides a source of benthic organisms for recolonization).  Nonetheless, the rate of temporal improvement in the debris-influenced deep areas of Quesnel Lake is expected to be slow in association with the slow rate of improvement in sediment quality, including the required organic carbon content.  Sediment deposition rate studies have indicated that the accumulation of 1 cm of new sediment in the Quesnel Lake deep areas will take six to seven years. INTEGRATED SUMMARY AND CONCLUSION The Mount Polley TSF breach had a direct physical and chemical impact on bottom habitat of Polley Lake and debris-influenced areas of Quesnel Lake.  Conventional sediment quality assessment indicated that concentrations of copper in sediments of these impacted areas were substantially elevated relative to baseline and/or reference and upper WSQGs based on strong acid digests in association with concentrations of several other metals that were elevated to lower magnitudes (arsenic, iron, and manganese).  Concentrations of TOC were also substantially reduced with the introduction of the inorganic debris.  Evaluations of sediment quality using a suite of digestions (that included less aggressive extractants) indicated that the majority of sediment-associated metal was present in forms that could only be mobilized by strong acid and/or acid and a strong oxidizer.  This fraction is not considered to be mobile in aquatic environments.  Results of the extractions were consistent with known mineralogy and with geochemical investigations of Mount Polley tailings (including specific investigations of spilled tailings).  In Polley Lake, evaluations of acid volatile sulphides and simultaneously extracted metals indicated that sulphides were in excess and that metals should be predominantly bound by the sulphides.  Quesnel Lake sediments differed substantially from Polley Lake sediments as both organic matter and acid volatile sulphides were present at very low concentrations.  Results of monitoring using DGT devices supported the low metal mobility suggested by the combination of tailings geochemistry, selective extraction analysis, and the evaluation of sediment binding constituents – DGT-labile metal concentrations were low and have been stable since the initiation of DGT monitoring in 2015.  Furthermore, DGT-labile copper concentrations were well below those at which toxicity has been observed in other studies.  Standard sediment toxicity tests (10-day tests of Chironomus survival and growth; 14-day tests of Hyalella survival and growth) indicated no unacceptable effects to survival or growth of either organism associated with Polley Lake sediment.  This is consistent with the geochemical observations of low metal mobility, with SEM-AVS theory, and with sediment TOC content.  In Quesnel Lake, effects to survival and growth of both organisms were observed, which is consistent with high metal concentrations, low AVS, and low TOC.  However, metal concentrations documented in the toxicity test vessels differed substantially from that of in-situ overlying waters (e.g., copper concentrations were 40-times higher in the toxicity test vessels than in-situ), bringing into question whether these test results can be extrapolated to in-situ conditions.  Nonetheless, this testing indicated that low TOC was a key factor in the observed effects (as TOC was well below levels known to result in adverse effects in tests, even when organisms are fed per standard methodology).  The benthic invertebrate community of Polley Lake was impacted by the breach debris, but improved rapidly and was similar to historical and reference by 2016, which is consistent with the geochemical findings and the sediment toxicity testing.  The benthic invertebrate community of the breach-influenced deep areas of Quesnel Lake was also impacted by the breach debris, and, although recovery has been documented with each year of monitoring, the community remained different than reference on the basis of lower density, lower taxon richness, and community composition.  This is consistent with the expected low rate of recovery in sediment quality, including the slow buildup of organic matter to levels that fully support aquatic life.  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