British Columbia Mine Reclamation Symposium

Site-specific water quality objectives for mine environmental management Bright, Doug; Bryant, Debbie; Eickhoff, Curtis 2014

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-1-  SITE-SPECIFIC WATER QUALITY OBJECTIVES FOR MINE ENVIRONMENTAL MANAGEMENT Doug Bright, Ph.D., R.P.Bio1, Debbie Bryant, M.Sc.1, and Curtis Eickhoff, Ph.D.2 1Environmental Risk Assessment,  Hemmera, Victoria, BC.  2Director - Ecotoxicology,  Maxxam Analytics, Burnaby, BC.  ABSTRACT Few environmental issues are of greater importance for the planning and approvals of new mines, operational management, and closure/remediation planning than the often complex hydrogeochemical interactions between surface waters that support aquatic life and mine wastes or recently disturbed areas. There is a strong reliance for permitting and approvals within British Columbia (BC) and other Canadian jurisdictions on generic provincial and national water quality guidelines to interpret aquatic ecological risk potential and protect freshwater life. The mining sector supports greater use of site-specific, risk-based water quality benchmarks, which necessarily require a more direct scientific understanding of aquatic ecological risks in watersheds of interest; however, the adoption of site-specific water quality objective (SSWQO) approaches has been limited. We discuss the methods used to develop generic water quality guidelines, including differences between Canadian (CCME) and BC water quality guideline derivation protocols. The relationship between generic guidelines and SSWQOs are discussed, along with the various practical approaches for development of SSWQOs based on toxicity data such as the water effects ratio (WER) method. Finally, we discuss the factors that may undermine the more widespread adoption of SSWQOs, and of the associated support for science-based decisions for environmental management in the mining sector. KEY WORDS Water quality, toxicity testing, risk assessment, site-specific, water effects ratio INTRODUCTION  Significant progress has been made over the last three decades towards minimizing the impacts of base metal, coal, and other types of mining on ecologically productive aquatic ecosystems. Such progress includes (i) improvements in both mineral extraction and effluent treatment technologies; (ii) an increased percentage of process water re-use; (iii) greater knowledge about the roots of acidic drainage and other forms of weathering/leaching from mine wastes (e.g., non-acidic drainage); (iv) mining design and management at landscape and watershed scales based on an improved understanding of site hydrology (e.g. for tailing management facilities and waste rock deposits); and (v) advances in our understanding of groundwater-surface water interactions, especially in the context of seepage down-gradient from waste deposits. At the same time, the average scale of proposed and new mines worldwide has increased by roughly an order of magnitude with each passing decade, as higher-grade deposits are depleted and as extraction technologies improve in efficiency and cost-effectiveness. Generic water quality guidelines have historically played a key role in mining environmental management, and will continue to do so in the future in the context of regulatory approvals, discharge permits, monitoring, and closure. According to the BC Ministry of Environment (MOE), an aquatic environmental impact assessment is “the main tool to inform the statutory decision maker of proposed risks of the project to the aquatic environment” (BC MOE 2013a), which in turn is informed by the comparison of data and predictions to the BC generic water quality guidelines or SSWQOs. Generic water quality guidelines, as promulgated by the BC MOE, the Canadian Council of Ministers of the Environment (CCME), the United State Environmental Protection Agency (US EPA), and other agencies are generally intended to “ensure protection of the most sensitive intended water use” (BC MOE 2013b). Use of SSWQOs as a management tool instead of generic water quality guidelines is not intended to alter the desired levels of protection of sensitive uses or aquatic ecosystem health, as defined through public and regulatory policy, but rather allow for greater consideration of the site-specific abiotic and biotic characteristics of lotic, lentic, estuarine, and marine ecosystems that could influence either the sensitivity of biota present or the degree of exposure relative to a measured concentration in water.  A common perception among some members of the regulated community about provincial and national water quality guidelines is that some of the values are overly conservative, leading to predictions of risks to freshwater life at waterborne trace element concentrations that would not impair the intended use of the water body to support the productivity and biodiversity of fish and other taxa of ecological, social, cultural, and economic importance. Perhaps it is for this reason that many mine environmental managers and mining associations such as the Mining Association of BC (MABC) and the Mining Association of Canada (MAC) have advocated for increased regulatory acceptance of the development and use of SSWQOs along with site-specific risk assessments in general.  This paper briefly discusses the derivation procedures used for development of generic BC and CCME water quality guidelines, along with the major reasons that a generic water quality guideline might not be an accurate predictor of a threshold of impairment to the most sensitive intended water use on a site-specific basis. We then summarize the four major approaches that have been prescribed by BC MOE and CCME for derivation of a SSWQO. These approaches are (i) the background concentration approach; (ii) the analytical limit of quantification approach; (iii) the re-calculation (resident species) approach; and (iv) the water effects ratio (WER) approach (CCME 2003, BC MOE 2013b). We provide several relevant examples in which such approaches have been applied, with a focus on the WER approach. Finally, we discuss potential impediments to the use of site-specific approaches for the mining sector in British Columbia, based in part on expectations about the level of effort required to develop a SSWQO relative to potential for (i) shifting environmental management objectives and (ii) regulatory and public acceptance.  This brief discussion is also based on the recognition that there is broad acceptance by both the environmental regulator and mining community of the value of environmental monitoring of aquatic effects as an important environmental protection tool, with a focus on field-collected population, community, and other biological data in surface waters that may be affected by current or past mining activities. Generic and site-specific water quality criteria, therefore, are only a subset of tools and approaches that are used for protecting aquatic resources and uses. ROLE AND LIMITATIONS OF GENERIC WATER QUALITY GUIDELINES  Contemporary derivation protocols for environmental quality guidelines begin with the collation of the existing available toxicological data that relate an adverse effect in an organism of interest and an exposure concentration or dose for a substance of interest. The ability to derive a meaningful and reasonably accurate (relative to narrative objectives) water quality guideline from the meta-analysis of relevant ecotoxicity data is limited by the availability and quality of the existing data. For aquatic ecotoxicity data, considerable effort has gone into developing internet-accessible and other biological effects databases. The fact that the existing primary data have typically been produced as part of various studies with objectives far removed from the identification of toxicological thresholds, however, suggests a need for a reasonable level of skepticism when attempting to define toxicity reference values intended to protect all multicellular species within aquatic ecosystems.   Elements of the derivation procedure that are common to the BC MOE (2012), CCME (2007), and the US EPA (1985) approaches include toxicity data collation, screening, and manipulation, as illustrated in Figure 1. The BC MOE approach departs from the US EPA approach and CCME (2007) revised approach in a few major respects. First and foremost is that the preferred CCME and US EPA option, the availability of data permitting, uses a species sensitivity distribution approach (SSD). The SSD approach amalgamates concentration-based toxicological thresholds for each aquatic species from the lowest to highest concentrations. For example, an estimated concentration from experimental (laboratory toxicity)  results associated with a 20% reduction in the growth or reproduction of a species (i.e., an EC20, or concentration associated with a 20% effect size) is tabulated, and the relative ranking of an EC20 value for each species in the tabulated data provides an indication of its relative sensitivity. The nomination of a value at or very close to the lower end of the SSD (e.g. the 5th percentile value) serves as a generic threshold that should adequately protect all potentially present aquatic species.   The degree of conservatism in achieving narrative protection goals for an SSD-type approach depends especially on three aspects of data collation and manipulation: (i) the specific types of toxicological Figure 1: Routine approach for the development of generic water quality criteria endpoints selected to develop the SSD; (ii) whether there are additional uncertainty factors (UFs) applied to the actual individual statistical estimates for each species and toxicological endpoint; and (iii) the point on the SSD that is selected as the generic water quality guideline, allowing for upper and lower confidence intervals in the statistical curve-fitting for the SSD.  CCME (2007) prescribes the use of toxicological data from available studies (or ECX estimates secondarily calculated from contributing studies) with the following order of preference:  EC10/IC10 > EC11-23/IC11-23 > MATC > NOEC > LOEC > EC24-49/IC24-49 > non-lethal EC50/IC50 IC stands for inhibitory concentration, while the maximum acceptable toxicant concentration (MATC) is calculated as the geometric mean of the no observed effect concentration (NOEC) and lowest observed effect concentration (LOEC).  The preference for ECX/ICX data, where X <<50%, and especially ≤10% raises some important issues, not the least of which is that the statistical robustness of the point estimate on the dose-response curve is low in comparison to an EC50 value, since it typically involves a statistical interpolation (or rarely an extrapolation) well beyond the observational range of most of the experimental data.  Furthermore, an EC10 estimate for most aquatic toxicity studies is rarely statistically, significantly different than the EC0; i.e., the response in the experimental controls. In fact, the CCME Type A approach is specifically intended to preferentially and primarily plot no-effect data for the derivation of SSDs (CCME 2007).   The vast majority of available toxicological data for aquatic life is also expressed as an acute or sub-acute LC50 (lethal concentration for 50% of the test organisms) or EC50, which is not a favoured data type in the CCME approach. While it can be argued that an ambient environmental concentration no greater than a chronic LC50 value would allow some members of the affected sub-population of the species of interest to survive, and would be commensurate with the persistence of all less sensitive taxa, neither BC MOE nor CCME favour this argument. In fact, narrative statements on aquatic environmental protection goals, as published by BC MOE and CCME, are sufficiently vague that there is no means of exploring what is or is not likely to be a significant aquatic ecological effect (or impairment of use for aquatic life) based on constructs such as ecological productivity, function, and biodiversity. CCME water quality guidelines are intended to “protect all species all the time” (CCME 2007). A key message, therefore, is that the development of SSWQOs will not be useful in circumstances where there is a fundamental disagreement with policies and practices on how the narrative protection goals are translated into generic numerical guidelines.  Under the BC MOE approach (Meays 2012), the preferred ranking of available toxicity data is similar to the CCME approach:  ECX/ICX representing low-effects threshold > EC15-25/IC15-25 > LOEC > EC26-49/IC26-49 > non-lethal EC50/IC50 > LC50 The development of a generic water quality guideline, however, is not based on an SSD approach. Rather, the lowest relevant toxicity endpoints are scrutinized and an uncertainty factor, typically in the range of 2 to 10, is applied to derive a short-term maximum and long-term average water quality guideline. Not surprisingly, one of the major points of contention for specific BC water quality guidelines is in how professional judgments have been applied in the nomination of uncertainty factors.   Under both the BC and CCME approaches, as formally described in their derivation documents, there appears to be some confusion about whether an EC effect size less than 10% to 15% is a de facto no effect or low effect level.  For recent CCME and BC MOE derivations of generic water quality guidelines, the role of exposure and toxicity-modifying factors (ETMFs) (CCME 2007) has received considerable attention. Such factors potentially include pH, alkalinity, hardness, dissolved oxygen (or redox), and temperature. Several generic water quality guidelines include specific predictive regression-based estimates of the threshold of toxicity to freshwater life based on relationships with pH or hardness. For the BC water quality guidelines, these include aluminum, cadmium (in draft), copper, fluoride, lead, silver, sulfate, and zinc. The absence of reference to ETMFs in currently adopted and working generic water quality guidelines does not mean that such modifying factors are not potentially important to an understanding of aquatic ecological risks, but rather that there likely has been inadequate scientific knowledge to incorporate adjustments for ETMFs into the derivation. In addition, the mathematical formulae that have been incorporated in the above-listed guidelines to account for ETMFs should not be viewed as over-arching scientific truths but rather as best approximations of toxicological potential from total waterborne concentrations and important co-variates based on the scientific data available from studies of a very limited subset of species that are of interest (under the premise that all models are wrong; however, some are useful).  The vast majority of derivation protocols for provincial, national, and international water quality guidelines have – by policy – attempted to minimize the probability of committing Type II errors (Figure 2) at the expense of inflating the probability of making Type I errors; i.e., predicting effects to aquatic life in the receiving environment based on reference to generic water quality guidelines when in fact no effects are present. As an environmental quality guideline approaches a value of zero, the probability of a Type II error approaches zero; however, the probability of a Type I error approaches 100%, with commensurate societal implications for the investment of resources in environmental protection beyond any real potential for environmental gains.    For the promulgation of generic environmental quality guidelines in general, the recognized degree of conservatism in the resulting environmental effects thresholds is often rationalized in part based on the fact that generic environmental quality guidelines are necessarily applicable across a broad range of ecosystems, biological species assemblages, habitats, and communities, and must be able to provide adequate protection for more sensitive versus average conditions. The use of site-specific approaches, therefore, should in theory allow a better balancing of the probability of making Type I versus Type II errors.  PRESCRIBED APPROACHES FOR DEVELOPMENT OF SSWQOS The supporting rationale and specific methods for development of SSWQOs are documented elsewhere (BC MOE, 2013b; CCME, 2003). Of particular note is that the BC MOE approach for SSWQO is prescriptively constrained, allowing for the application of one or more of four possible approaches. Each of these is briefly described below from the perspective of regulated users.   Background Concentration Approach It is not reasonable to manage human activities toward the achievement of chemistry-based numerical objectives in the receiving environment that are lower than naturally occurring background concentrations; this principle is broadly recognized in the development of provincial and national regulatory policy and guidance. In the context of mining, it is common to observe the concentrations of some trace elements dissolved in water, in suspended sediments, or in bed sediments at concentrations that exceed the generic guideline for aquatic life protection, For such circumstances, BC MOE prescribes adopting as the preliminary SSWQO an upper limit of natural background concentration based on the 95th percentile concentration.   The Background Concentration Approach is useful in a minority of circumstances in BC for mining environmental management. We have generally found within our studies that the initial perceptions about naturally elevated trace elements are greater than the reality, particularly for downstream areas of watersheds that broadly integrate meteoric water from both highly mineralized areas with surface manifestations and adjacent non-mineralized sub-watersheds. Furthermore, the approach is generally relevant to pristine watersheds, and it becomes more challenging to define through field studies an Predictions of Effects to  Aquatic Life From Comparison of Discharge of Ambient Concentrations to WQGActual Effects to  Aquatic LifeNo Effect EffectEffectNo Effect CORRECT PREDICTIONCORRECT PREDICTIONTYPE IERRORTYPE IIERRORFigure 2: Type II versus type I error in the application of water quality guidelines appropriate upper limit of natural background concentrations in sub-watersheds that may have already been influenced by exploration and extraction activities.   The Background Concentration Approach is sometimes useful as a quick fix for BC working and approved generic water quality guidelines, or CCME guidelines, that were derived using older and different methodologies, and under an era when peer review was not as important in the overall process. The boron BC water quality guideline for protection of marine life, updated in 2003, was set at 1,200 µg/L even though the guideline derivation document states “median values for boron in surface water is about 0.1 mg/L and in Canadian coastal marine water it ranges from 3.7 to 4.3 mg/L”(Moss and Nagpal 2003).   Analytical Limit of Quantification Approach Rarely, a generic environmental quality guideline has been developed that is lower than the analytical quantification limit based on reasonably accessible laboratory services, methods, and instruments. This occurs for one of the following reasons:  1. The provisional generic water quality guideline was derived through the application of uncertainty factors or other lower-concentration extrapolation procedures that resulted in a calculated threshold for protention of aquatic life protection that is substantially lower than any concentration referenced in any of the underlying laboratory ecotoxicity studies.  2. The critical ecotoxicity study(ies) that strongly influence the provisional water quality guideline did not use a measured exposure concentration, but rather assumed a spiked (nominal) exposure level (e.g., based on serial dilutions), often without due consideration of the associated uncertainty.  3. The researchers that carried out the critical ecotoxicity studies had access to highly specialized, research-quality chemical analytical tools that are generally not readily available beyond specialized research institutions.  4. The water quality guideline is based on back-calculation from a tissue residue concentration, wherein the critical threshold has been based on bioconcentration or biomagnification.  Modern water quality guideline derivations should adequately account for such possibilities; however, water quality guidelines derived more than a decade ago may require adjustments to reflect practical limits of quantification. According to BC MOE (2013b), “A provisional WQO could be set at the achievable analytical level until the improved detection limit can be established.”  Recalculation Approach (Resident Species Approach) The USEPA (2013), CCME (2007), and BC MOE (2013b) allow for the re-calculation of water quality guidelines from the available credible aquatic toxicity data used in the original derivation of the generic water quality guideline to incorporate data for only those taxa that are potentially present in the water bodies of interest. The particulars of the prescriptive guidance for each authority are different but the underlying principles and procedures are the same: The aquatic biota that are present or potentially present within the water bodies of interest need to be defined through references to authoritative sources and based on field surveys. The list of actual potentially present taxa is then cross-referenced against the collated ecotoxicity data that underpin the derivation of the generic environmental quality guideline: Taxa within the ecotoxicity database that are not relevant to waters of interest are omitted and the water quality guideline is re-calculated based on the prescribed procedures for the derivation of generic water quality guidelines.  In practice, this site-specific species selection of recalculation approach has limited utility. Since ecological risks assessments broadly rely on data from surrogate species, it is often difficult to eliminate specific data for one or more potential surrogate species as being irrelevant to the species that might potentially inhabit the site of interest. Furthermore, when assessing trace element, major ion, or nutrient toxicity in surface waters  the most sensitive toxicity endpoint in the database is often from a chronic (e.g., reproductive impairment) test on a cladoceran such as Ceridaphnia dubia  or  Daphnia magna, for which allied species within the same order and family are very widely distributed in temperate to arctic surface waters. In addition, the elimination of aquatic ecotoxicity data for absent species can have the effect of imposing further data limitations. Checks and balances, as part of the recalculation approach, include the assurance that sufficient ecotoxicity data remain in the reduced database that are of relevance to the site-specific list of aquatic organisms. It may be necessary, therefore, to undertake additional laboratory toxicity studies to meet minimum data requirements for development of a SSWQOs. Under the BC MOE approach, the smaller the available ecotoxicity data set, the larger the uncertainty factor that is likely to be accepted in defining a SSWQO.  Water Effects Ratio Approach The Water Effect Ratio (WER) procedure is useful in circumstances in which factors related to the physical and chemical characteristics of the mine discharge receiving water are expected to modify the toxicity of the trace element(s) of concern. For example, high levels of total suspended solids (TSS), dissolved organic carbon (DOC), or total organic carbon (TOC) may indicate the presence of high concentrations of organic ligands such as humic and fulvic acids that may interact with metals/metalloids and alter their bioavailability and/or toxicity in the water column. If the receiving water has high concentrations of organics, the WER approach is likely to provide a SSWQO for trace elements such as copper or zinc that is higher than the generic water quality guideline. Adequate knowledge about the aquatic geochemistry of the substances of interest will greatly assist with decisions about whether a WER approach is potentially helpful. For inorganic arsenicals, the scientific literature is somewhat equivocal but nonetheless indicates that organic matter is not a strong sorptive media or complexing agent in comparison with iron and manganese oxyhydroxides. A WER approach, therefore, is likely to have limited value in producing a SSWQO that is substantially different than the generic water quality guideline for inorganic arsenic.  An advantage of the WER approach is that it directly compares the actual toxicity of a substance of concern – across a range of concentrations – in the receiving water and in standardized laboratory test water. Aquatic toxicity bioassays typically are performed with pristine laboratory control or dilution water. Generally, the water is treated by filtration and TOC is normally very low. Many of the aquatic toxicity bioassays such as acute Daphnia, and sublethal tests with algae (Pseudokirschneriella subcapitata), invertebrates (Ceriodaphnia dubia), and aquatic vascular plants (Lemna minor) are conducted with reconstituted waters prepared from purified deionised water. These recipes, specified by the test guidelines, generally have very low TOC content; therefore, toxicity tests conducted on an effluent with pristine lab water as a diluent may produce different results than tests conducted using receiving water that is high in TOC.  It is recommended that bioassays are conducted with at least two indicator species using site or receiving water and lab water (see Table A5.2 of BC MOE, 2013b). Typically, an acute and a sublethal or short-term chronic toxicity test will be conducted with organisms that are chosen based on the applicability to the system, availability, and sensitivities to metals and metalloids in freshwater systems. The 96-hour rainbow trout and 48-hour Daphnia magna are commonly used to assess acute toxicity. Fathead minnows (Pimephales promelas), water fleas (C.dubia), freshwater algae, and other species may be used to evaluate longer-term effects to growth and/or reproduction. Other species may be used that are considered resident or particularly sensitive to the substance(s) of concern.   For metals, a stock solution is prepared using a readily soluble salt (e.g.  nitrate, chloride, or sulphate salt) that is spiked into both lab water and site water at the same range of concentrations. A range-finding study that investigates a wide range of concentrations is helpful to better narrow the concentration range of interest for subsequent definitive tests (i.e., to narrowly bracket the toxicological endpoint of interest). The concentration of the metal (dissolved and total) should be determined in the test solutions at the beginning and end of the tests to ensure that the measured concentrations are similar to the nominal or expected concentrations. This determination is essential to understanding the results of the tests and to ensuring the quality and accuracy of the test endpoint (LC50) values.  In some instances, obtaining the correct metal concentrations in the test solutions may be challenging if the metals interact with elements of the lab or site waters. For example, carbonate in the recipe of the Ceriodaphnia dubia test water may react and precipitate out the test metals, which may reduce the expected concentrations of metal available in laboratory water solutions; this will reduce the concentration of dissolved metal in the tests over time and may result in reduced toxicity in the laboratory test water.  In site waters, calcite super-saturation can have a similar influence. Other quality assurance and quality controls are incorporated to identify deviations from the test procedures, and problems associated with temperature, loading rates, feeding, dissolved oxygen, pH, control failures, and other issues that may compromise the reliability of the bioassay results.   The WER value is calculated for each species based on the results of the bioassays conducted using the spiked site/receiving water and laboratory water. Data from the toxicity tests are used to calculate the appropriate statistically derived endpoints such as the LC50, EC50, or IC50 values. The WER value is calculated as follows in Equation 1:   WER = Site Water LC50/Lab Water LC50    [1] If the WER values for the two species are similar, (within three-fold of each other), then the geometric mean of the two values is calculated and used to modify the generic BC WQG as shown in Equation 2:    SSWQO = WQG x WER      [2] If the two WER values are not similar, then an additional set of paired tests is conducted with another relevant species to confirm or refute results of the two initial tests. The SSWQO would then be derived by multiplying the BC generic water quality guideline by the WER calculated as the geometric mean of the two similar lowest WER values. For metals, the final WER is calculated for both total and dissolved concentrations of the metal of interest.  EXAMPLE OF THE WER APPROACH FOR SSWQO ESTIMATION:  Maxxam performed a WER experiment to support development of an SSWQO for copper for a mine project in the BC and Yukon region. Aquatic toxicity tests were performed on two species, using both acute and chronic exposure durations: 96-h rainbow trout acute test (per Environment Canada EPS 1/RM/13); Ceriodaphnia dubia 48-h acute test, and 7-d, three-brood reproduction test (per the US EPA 2002 acute method and Environment Canada 2007 chronic method).   Total and dissolved metal concentrations, water hardness, TSS, and TOC were measured in the site water and the hardened laboratory water. The site water total hardness calculated based on the metal analysis was on average 200 mg CaCO3/L. The TOC was 2.47 mg/L and the TSS was below 1 mg/L. The laboratory water hardness used for the fish tests was adjusted to 204 - 212 mg CaCO3/L to match the site water. The TOC was 0.66 mg/L and the TSS was below 4 mg/L. For the Ceriodaphnia tests, the laboratory dilution water was made with Type I deionized water adjusted to 188 mg CaCO3/L hardness with Perrier water, reflecting the hardness of the site water, and addition of vitamin B12 & selenium to maintain the health of the organisms. The culture water used was made with deionized water hardened to 96 mg/L CaCO3 with 20% Perrier water, vitamin B12, and selenium.  Stock solutions were prepared using a highly soluble reagent-grade CuCl2 spiked separately into the laboratory water and the site water. The nominal copper concentration ranges were 0 µg/L to 228 µg/L in laboratory water, and 0 µg/L to 638 µg/L in the site water. Prior to initiating the tests, subsamples were collected from all test concentrations for the analysis of total and dissolved copper. All test solutions were further subsampled for analysis prior to and after a 48-hour water renewal and at the end of the tests.   To calculate the WER value for copper, the laboratory and site water 96-hour LC50 were estimated by probit analysis based on measured copper concentrations of the test solutions. The subsequent WER values, calculated based on measured concentrations of copper in test solutions, are presented in Table 1.  Table 1: Results of Water Effects Ratio Testing Test LC50 Lab. Water  (Diss. copper µg/L) LC50 Site water (Diss. Copper µg/L) WER  Acute rainbow trout 115.7 (97.64-137.1) 188.2 (156.5-225.9) 1.6 Acute Ceriodaphnia dubia  3.2 (2.9 - 3.4) 12.0 (9.1 - 16.7) 3.8 Chronic C. dubia survival  6.8 (5.7 - 8.1) 33.8 (30.3 - 37.7) 5.0 Chronic C. dubia  reproduction 3.6 (2.8 - 4.2) 23.7 (17.8 - 28.3) 6.6  The final WER values for copper for the mine site in question were obtained by determining the geometric mean of the comparable WER values obtained from the toxicity tests. In this instance, the WER values obtained from the chronic Ceriodaphnia tests were higher than that obtained for the acute trout and Ceriodaphnia tests. The guidance for deriving site-specific water quality objectives (BC  MOE 2013b) states that when more than two WERs have been determined, the final WER should be calculated as the geometric mean of two lowest WER values. Therefore, in this study, the final WER value was calculated as the geometric mean of the WER values obtained from the acute fish and Ceriodaphnia tests as follows:   Final WER = geometric mean (1.6, 3.8) = 2.5     [3] In this example, the water quality objective for copper was then calculated based on the water quality guideline for copper. The water quality objectives for the protection of aquatic life for copper obtained from BC MOE are presented in Table 2.  Table 2: Summary of Water Quality Criteria for Copper Criterion 30-day Average µg/L Total Copper Maximum µg/L Total Copper Freshwater Aquatic Life  (when average water hardness as CaCO3 is greater than 50 mg/L) <0.04 (mean hardness) µg/L (0.094(hardness)+2) µg/L (hardness as mg/L CaCO3) WQG  8.0 21 SSWQO 20 51  Based on a mean site water hardness of 200 mgCaCO3/L, the water quality guideline for copper would be <8 µg/L Total Copper for the 30-day Average and 21 µg/L Total Copper as a maximum concentration. The SSWQOs were then calculated based on Equation 2 and are presented in Table 2. The WQG values were multiplied by the WER value of 2.5, which resulted in SSWQO values of 20 µg/L and 51 µg/L total copper for the 30-day average and maximum concentrations respectively.   CONCLUSIONS  The SSWQO derivation approaches prescribed by BC MOE and CCME are of practical value to mine environmental planners and managers under specific circumstances. As discussed above, three of the four prescribed approaches (background concentration approach, limit of quantification approach, recalculation approach) are likely to be relevant to only a very limited number of substances and site-specific receiving environments. The water effects ratio approach has proven to have good practical value for sites where cadmium, copper, lead, silver, and zinc are substances of interest (Nautilus Environmental 2009; Diamond et al. 1997; Jop et al. 1995). The WER approach has been used for copper more than any other trace element. An obvious challenge with the WER approach is that it can provide only limited compensation for a generic water quality guideline that is based on potentially unrepresentative primary data, or is unduly influenced by conservatism in the derivation approach, since the SSWQO is a product of the generic water quality objective and the WER.  The available approaches for development of SSWQOs do not address several aspects associated with application of generic water quality guidelines. Among these are cases where the existing approved and working generic water quality guideline do not provide an accurate estimate of the threshold of toxicity based on (i) non-reproducible primary data that drive the lower end of the SSD or are among the lowest values available at the time of derivation; (ii) use of uncertainty factors or other compensatory approaches in the face of data limitations; (iii) estimation of EC10 or similar endpoints from dose-response data with the attendant degree of statistical uncertainty; and/or (iv) the chemical form of the substance of concern in the site receiving environment relative to the readily soluble metal salts that are typically used in WER studies. Often, the best approach for dealing with scientific or technical limitations in the existing generic water quality guidelines is to complete focused scientific studies to address knowledge gaps (e.g., to better understand modifying factors for exposure and toxicity) that are not accommodated within the existing prescribed SSWQO approaches. Furthermore, the prescribed SSWQO approaches cannot reconcile philosophical differences in professional opinion about the translation of vague narrative protection goals into concrete management objectives, or resolve perceptions regarding the effectiveness of aquatic environmental protection  on a case-by-case basis as a result of interpretations of water chemistry relative to water quality guidelines versus biological survey data on resident species, communities, biodiversity, and productivity.  REFERENCES  British Columbia Ministry of Environment (BC MOE). 2013a. The effluent permitting process under the Environmental Management Act:  An overview for mine project applicants. 19 pp. (accessed July 19, 2014). BC MOE. 2013b. Guidance for the derivation and application of water quality objectives in British Columbia. 147 pp. (accessed July 19, 2014).  Canadian Council of Ministers of the Environment (CCME). 2003. Guidance on the site-specific application of water quality guidelines in Canada: Procedures for deriving numerical water quality objectives.,d.cGE (accessed July 19, 2014).  CCME. 2007. A protocol for the derivation of water quality guidelines for the protection of aquatic life 2007. 37 pp.,d.cGE (accessed July 19, 2014).  Diamond, J.M., Koplish, D.E. McMahon J. III, and Rost R. 1997. Evaluation of the W\water-effect ratio procedure for metals in a riverine system. Environmental Toxicology and Chemistry 3:509–520.  Environment Canada. 2000. Biological test method: Reference method for determining acute lethality of effluents to rainbow trout. Report EPS 1/RM/13. Ottawa, ON.  Environment Canada. 2007 Biological test method: Test of reproduction and survival using the cladoceran Ceriodaphnia dubia. Report EPS1/RM/21 Ottawa, ON.  Johnson T. R., 1987. Water quality criteria for copper-Overview report. BC Ministry of Environment, Environmental Protection Division. (accessed July 19, 2014).  Jop, K.M., Askew, A.M., Foster, R.B., 1995. Development of a water-effect ratio for copper, cadmium, and lead for the Great Works River in Maine using Ceriodaphnia dubia and Salvelinus fontinalis. Bulletin of Environmental Contamination and Toxicology 54: 29-35.  Meays, C., 2012. Derivation of water quality guidelines to protect aquatic life in British Columbia. 34 pp. (accessed July 19, 2014).   Moss, S.A and N. Nagpal, 2003. Ambient water quality guidelines for boron: Overview. (ccessed July 20, 2014).  Nautilus Environmental, 2009. Water effects ratio for Minto Mine–copper. 16 pp. (accessed July 20, 2014). 


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