Life-Cycle Inventory Procedures for Long-Term Release of Metals

inventory issues in LCA, and review existing and proposed approaches to make LCA ... This implies that exposure models and inventory methods both must...
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Critical Review Critical Review: Life-Cycle Inventory Procedures for Long-Term Release of Metals JOHAN PETTERSEN* AND EDGAR G. HERTWICH Industrial Ecology Programme and Department of Energy and Process Engineering, Norwegian University of Science and Technology (NTNU), NO-7491 Trondheim, Norway

Received August 29, 2007. Revised manuscript received March 25, 2008. Accepted April 7, 2008.

Life-cycle assessment (LCA) is the method of inventorying, assessing, and interpreting environmental interventions caused by products and product systems through their life cycle. The ecotoxicity of metals has proven a challenge for LCA given metal characteristics such as reversibility of removal processes, speciation, and the effect on bioavailability and ecotoxic effect assessment. Our review focuses on the first part of the ecotoxic impact chain for metals, i.e., the release of metals from solid deposits. According to the principle of temporal justice, sustainability assessment tools such as LCA should account for emissions regardless of temporal location distribution. This is in LCA commonly interpreted as leaching until depletion of metals bound in solid wastes under the presumption that infinite time implies infinite weathering. This approach is risk conservative for metals and it hampers the use of LCA to assess remediation projects for soils and sediments contaminated by inorganic substances. We discuss metal significance and inventory issues in LCA, and review existing and proposed approaches to make LCA applicable to metal long-term emission.

Introduction Life-cycle assessment (LCA) aims to provide a tradeoff of environmental impacts both along the life cyclesfrom cradleto-gravesand across different impact categories. In LCA, emissions from product systems are hence inventoried, assessed, and interpreted (1–4). Emissions are translated to environmental impact scores in the impact assessment stage of LCA. Methods for assessment of ecotoxic impacts have a long tradition in LCA and a number of approaches have been developed to evaluate the environmental distribution, fate, and effect of ecotoxic substances (5). Metals have been found to dominate ecotoxic impact category results for a variety of product cycles (6–10). Existing impact assessment approaches rank metals among the most ecotoxic both in terms of their no-effect thresholds and time-integrated toxicity (11–13). In this paper we review the treatment of long-term metal emissions in LCA. We address problems and uncertainty, solutions in related fields of risk assessment and waste characterization, and implications for decision support. Metals have proven to be a challenging class of substances in risk and impact assessment. Challenges related to metal reversible immobilization and removal processes, aquatic phase speciation, various uptake routes, and potential * Corresponding author phone: (+47) 735 989 55; fax: (+47) 735 989 43; e-mail: [email protected]. 10.1021/es702170v CCC: $40.75

Published on Web 05/21/2008

 2008 American Chemical Society

essentiality are common to environmental geochemistry (14, 15), risk assessment (16–18), and life-cycle assessment (19, 20). The complexity of metal fate in the environment affects life-cycle ecotoxic assessment in three ways. First, it complicates the estimation of metal emissions from solid deposits. Typical examples include leaching from mine tailings and landfills (21–24), but contrary to most organic emissions metals are released to nature in a number of physical-chemical states. Long-term emission estimates require some model definitions which carry significant uncertainty (25, 26) even if emissions may be extrapolated for a finite period from laboratory or site measurements with good precision. It may be argued that an extended time perspective allows for the completion of weathering, even if occurring at slow rates. However, there are simultaneous processes working to stabilize metal in solid states (27–30). Second, exposure depends on metal speciation, which again is site dependent. Local environmental conditions also affect bioavailability and the reversible immobilization of metals. This implies that exposure models and inventory methods both must be spatially resolved (16, 19, 31). This is a substantial problem for life-cycle assessment as the various processes that make up a product system may be globally distributed, and the locality of emission points is largely uncertain. To assume that the various emissions in the end may be approximated by an average situation, which is a valid approach for predominantly distributed emissions such as emissions to air from transport operations, may not hold for metal emissions as they often are dominated by large point-sources leaching metal from solid wastes; see for instance a recent study of global emission inventories for silver (32). The aspect of local condition uncertainty therefore is significant. Third, ecotoxic indicators should be representative of ecosystem response. This relates to LCA as a tool for decision analysis and hence the demand for operational attributes (33). Depending on local environmental conditions, various uptake pathways may become relevant (34, 35). A limited definition of ecotoxic pathways in LCA may therefore not properly represent the ecotoxic stress posed by metal emissions. An overview of the issues relating to ecotoxic assessment of metals in LCA is given in Figure 1. In this review we focus on the following points: (1) the significance of solid wastes to LCA results and uncertainty in long-term metal emissions from solid wastes; and (2) methods to treat this uncertainty when LCA is used as decision support, including proposed and implemented solutions in LCA, as well as possible approaches used in related fields. VOL. 42, NO. 13, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Processes in the metal ecotoxic impact chain, their characteristics, and relation to LCA. Based on Smith and Huyck (36). Before we start it is useful to define some of the terms used in Figure 1 and in the remainder of this paper. We adopt the terms geoavailability, mobility, and dispersivity from Smith and Huyck (36). Geoavailable metal is the portion of total metal that, within the limits of mobility and dispersivity, can be made bioavailable. Mobility describes the physio-chemical processes that control aquatic and solid state metal speciation, phase distribution, and reactive potential. Dispersivity refers to the physical processes that drive distribution by nonchemical means. Examples of dispersivity are percolation, sediment transport, erosion, and wind. Bioavailability refers to availability for biological uptake, i.e., the degree to which a contaminant is free for uptake (37), analogous to the common LCA terminology (19). Leaching is a term most often used with reference to laboratory procedures. By leaching we intend the use of an extractant to transfer metal from a solid sample. Leachability is the degree to which a metal may be released by leaching.

Metal Significance and Inventory Issues Life-cycle assessment has identified emission of metals as major ecotoxic contributions from a range of products and product systems (6–10). Moreover, existing impact assessment approaches rank metals among the most ecotoxic both in terms of no-effect thresholds and time-integrated toxicity (11–13). The latter are characterization factors in the terminology of LCA. Characterization factor are quantitative scores for the relative ecotoxicity of substances to defined environmental recipients through set pathways of exposure (4). Example Case: Diesel Energy. As an illustration of the significance of metal leaching from solid wastes in LCA we define an example function: energy supplied by diesel

generators operating in an average European situation (38). Conclusions from this simple study carry relevance for many LCAs given that processes and emissions invoked by energy use tend to be important for a variety of product systems. Huijbregts et al. (11) are the only source of ecotoxic characterization factors for the marine environment and are therefore used throughout this paper. Ecotoxic contributions from diesel energy are illustrated in Figure 2. We may conclude from the results that inorganic emissions in this case dominate ecotoxicity. This should motivate precision in emission estimates for metals and other inorganics. Moreover, solid waste fractions contribute significantly to all three recipients, underlining the importance of long-term emissions to the results. Finally, along with LCA practice we have aggregated direct emissions, such as produced water, with potential long-term emissions estimated from total metal contents in solid wastes. As a result, barite (BaSO4) discharges during offshore drilling operations dominate marine aquatic ecotoxicity although the potential release of barium from barite is considered miniscule compared to the total content (39). LCA vs Risk Assessment. The goal of LCA is to aggregate and assess environmental interventions independent of when and where they occur. Main guidelines allow limitations in the time frame that is applied for emissions (3), but the LCA community maintains that life-cycle assessment is a tool for sustainability assessment and therefore must incorporate the principle of temporal justice. Emissions in life-cycle inventories therefore should be aggregated until infinity in LCA (40, 41). Current impact assessment methods do not separate on temporal location. Time-integrated emissions are therefore treated as if they occur simultaneously. Spatial resolution has been implemented for some impact chains to take into account site-condition variability (though not for ecotoxicity), but similar efforts to achieve temporal resolution are lacking (5). Several factors of importance to metal availability and mobility show large spatial variation including soil acidity (18, 42–44), oxidization-reduction (redox) potential, soil texture, clay and organic matter content, presence of Fe and Mn oxides, and additional cations (44–46). In the case that spatial variation is implemented, order of magnitude deviations may still remain between current state-of-the art model results (with large data requirements), and actual concentrations (47, 48). Problems remain with mapping of background metal concentrations onto which anthropological ecotoxic emissions are added (49), e.g., through critical load approaches (50). Seasonal changes and the potential for various future scenarios add to the complexity of estimating metal release and effect. Whether or not characterization factors are spatially resolved, challenges with locating emission points lead to

FIGURE 2. Ecotoxic contributions from diesel energy. The term Remainder refers to the sum of processes and substances each less than 5% of the total potential. Processes in grey refer to leaching from solid wastes. 4640

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uncertainty in impact assessment in LCA. This has been pointed out as a major issue for implementation of metal characteristics in life-cycle impact assessment (19), and represents an important difference between ecological risk assessment (ERA) and assessment of ecotoxicity in LCA. Ecological risk assessment, which is fundamentally timeand site-specific, has a stronger recipient focus but is also less comprehensive than LCA (51). Due to emissions in ERA being reported in the format of expected exposure rather than time-integrated mass flows, the issue of estimating infinite leaching potential is specific to LCA. Risk Conservative Bias. The dominance of metals to ecotoxicity in LCA may be a fair judgment of the potential to cause adverse effects, but it is often caused by lack of incorporation of waste and deposit characteristics in inventory procedures. A simple example is highlighted by Kosson et al. for the release of lead from cementitious synthetic waste. The release at pH 9 is 5 orders of magnitude less than the total lead content (52). Clearly, environmental parameters are critical for the evaluation of the potential ecotoxicity of such wastes, and leachate pH at the site of deposit will change the outcome of an LCA study. This variation dwarfs the generic factors proposed for inventory uncertainty in LCA (53). Long-term emission estimates are a product of measurements at current deposit sites, laboratory- to deposit-scale experiments, and/or the use of modeling tools (26, 54). All of these methods rely on projection into an unknown future of leaching rates and/or of environmental conditions. Estimates not found by measurement are naturally connected to larger uncertainty compared to measurements of immediate emission processes. Leaching emissions are therefore frequently labeled as potential rather than actual releases. Given the vast possibilities for dispersion and mobilization that infinite time implies, it may be tempting to set the infinite leaching potential equal to the total mass. This may be expressed directly, see, e.g., ref (26), or indirectly by extrapolating leaching until heavy metals are depleted, as in refs (22, 23). Besides the uncertainty aspect, estimates made by assuming complete release therefore are potential in the sense that they carry a risk conservative bias. This bias is also stated in the literature. For instance, leaching tests are criticized for representing only a fraction of the total content (55), although this is the original purpose of performing such tests. Others argue that leachability is an operational definition, and that the outcome is controlled by the procedure (25). Except for a temporal separation, still with the infinite potential defined as the total mass, no alternative approaches have been proposed. The risk conservative approach of systematically overestimating long-term emissions makes impact assessment irrelevant in some cases. Due to the marine environment being the final sink for waterborne emissions and the extended marine residence time of inorganic substances (56, 57), results for the marine ecotoxic impact chain are particularly sensitive to imprecise leaching inventories. There is a growing literature on immobilization and mobility of metals once released into the environment. Mobile metals may be defined as the fraction of metals in a solid phase that is exchangeable through ionic exchange. The metals bound in soil and sediment may thereby be separated into a labile (i.e., mobile or bioavailable) and a fixed (i.e., nonmobile, nonbioavailable) portion; i.e., by use of isotopic techniques (30). Several studies investigate changes in the mobility of metals in soils, showing that there is a degree of fixation, albeit occurring at slow rates (27–30). This attenuation may be explained by micropore sorption and/or changes in metal oxidization states. In any case, given

that attenuation may occur, it is not evident that for a contaminated solid material all will be mobilized by dissolution. While sorption and desorption processes may be completed within relatively short time periods, redistribution of metal oxidation states in most cases is a long-term process. The net effect of changes in redox conditions is hard to predict (58). There are indications that Fe/Mn nodule formation in soil may occur through repeated redox cycling (58, 59) in a process similar to the selective redox cycling seen in the ocean (60, 61). At the same time, both anoxic and oxic conditions may increase leachability by supplying dissolved organic ligands (62, 63). Metal Remediation. A benefit of distinguishing the mobility of metals is that it allows evaluation of remediation techniques. Although risk conservative for metal ecotoxicity, the current framework for inventory estimation may be considered sufficient for the purpose of evaluating life-cycle metal ecotoxicity of production systems if a separation is made between short-term and long-term leaching potential. However, it is not applicable to remediation processes if the long-term leaching potential is set equal to the total metal content. In contrast to degradable organic pollutants, inorganic material cannot be degraded. Remediation of metal-contaminated soils and sediments therefore consists of attenuation measures (64). Risk reducing as they may be, they do not count as positive interventions in LCA. This paradox can only be solved by accepting that metals are bound in solids at various levels of stability and that only parts of the total metal may be mobilized to bioavailable states. Given that a strongpoint of LCA is the ability to discover system trade-offs, LCA should be an ideal tool for evaluation of the net benefit of remediation alternatives. However, the literature on LCA and remediation is rather scant (65), and none consider long-term emissions from the treated soil as part of the scope (65–68). It is reasonable to assume that this is at least partly due to problems with implementation of mobility measures in LCA.

Methods to Assess Long-Term Emissions in LCA The LCA community considers LCA a tool within the portfolio of sustainability assessment (69). As a consequence, the principle of temporal justice applies to the consideration of present and future environmental and human recipients (40, 41). Emissions from solid waste materials may continue for a long time, very possibly beyond the operating time of a landfill site or other infrastructures. Aggregation of direct, measurable emissions with those of long time processes, such as the leaching from solid wastes, implies summing of emissions with large differences in the uncertainty. In order to make LCA applicable for policy support, various approaches have been proposed. The Surveyable Period. One solution to make LCA more policy-relevant with respect to long-term emissions is to introduce a separation between near-term emissions and long-term to infinite emissions (26, 41). The short-term period is often referred to as the surveyable period, set at 100 years or similar in the literature (25, 26) usually interpreted as the follow-up or operating time of a deposit or landfill site. Leaching tests combined with models for the release from solid mineral phases offer estimations with order of magnitude precision for the short-term period (48). Separating a surveyable period we may either choose to neglect emissions beyond this stage or have them aggregated and interpreted separately. Alternatively, the ecotoxic effect of metals may be viewed as a separate impact category altogether due to the difficulties with assessing metal fate and ecotoxic effect (19). VOL. 42, NO. 13, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Boundaries for the Infinite. Long-term release potentials beyond the surveyable period may be based on mass, i.e., the total mass initially placed in the landfill; e.g. ref (8), or time. In the latter case, all releases may be neglected if occurring beyond 100 years (70), or if they occur after the expected time of the next ice-age (21). Alternatively, a critical time period may be introduced at which all significant effects are completed (26). A variation of temporal cutoff is applied for some of the mining processes in the ecoinvent database, namely to neglect leaching potentials completely by assuming them to occur in the extremely far future (24). (The delay in this particular case is caused by sulfide precipitation.) A less discussed approach is to base the cutoff for long-term releases on the resulting exposure concentrations (25). Assuming continuously decreasing leachate concentrations, releases occurring after a set threshold leachate concentration is achieved is then excluded from the emissions inventory. Possible sources of threshold levels include inert waste criteria, background concentrations, and effect-based thresholds. Discounting. Discount rates reduce the significance of impact from emissions that occur in the less certain future compared to emissions that occur today. Arguably making LCA results more compatible with economic evaluation tools, discount rates introduce a weighting between future and present impacts and violate the scientific approach asked for in the ISO standard (3) and by many in the LCA community; e.g., refs (25, 40, 71). If discounting is applied, negative as well as positive discount rates may be relevant and their consequence on the outcome of a study should be evaluated by scenario analysis (72). Any scheme involving a temporal boundary for emissions or discounting of future emissions is arbitrary. While they may simplify the estimation of long-term emissions, they are inherently based on subjective valuations of future potential environmental impacts and therefore are not suited for generic implementation in LCA. Waste Leaching Standards. The long-term leaching of metals is an issue of concern in the treatment and use of solid wastes. Decision support for waste treatment and resource management policy relies on the use of standard tests. Kosson et al. outline a management scheme for implementation of waste leaching tests (52). An impressive amount of test data has been compiled that complements the management scheme, covering various waste fractions and test conditions (54, 73, 74). Besides the obvious value as a life-cycle inventory source, conclusions made regarding the long-term leaching potential of solid wastes are important to the LCA community. Contrary to a general assumption of complete release, waste leaching takes as a starting point that three levels of leaching potentials may be identified (54, 73): (1) the total mass placed in the landfill; (2) the potentially leachable fraction; and (3) the fraction that is actually released under given environmental conditions. Tests to assess leaching may be classified as scenario analogy tests, sequential extraction schemes, and fundamental leaching parameter investigations (52). Scenario analogy tests are used in waste management generally as pass/fail criteria. Designs therefore are simplistic and risk conservative. Often referred to as “availability tests” (25, 52), operationally defined scenario tests include both single extractant schemes and consecutive acid extractions; e.g., refs (75, 76). Criticism has been raised against the use of availability tests for quantification of leaching in LCA, arguing that results are not consistent with the infinite leaching potential (25). Their use in LCA has also been proposed at the impact assessment level rather than within inventory estimation (77). Test observations confirm that consecutive extraction with pH 7 and pH 4 usually result in overestimation of the 4642

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FIGURE 3. Leachable metal per kg waste plotted as function of (a) pH, and (b) percolation volume. Redrawn from ref (54) to a generalized situation. Condition variability includes controlling parameters such as presence of organic phases, redox conditions, etc. release of metals compared to realistic environmental conditions with respect to percolation rate and pHsat least when not taking into account the influence of changes in redox conditions. Guidelines state pH within 3-5 to be the lower limit of what may be observed in nature (78, 79). While most environments have higher pH levels, availability tests therefore include extractions at pH 4. Leachability generally is lowest at near-neutral pH, meaning that acid extractions give high estimates. The conclusion of risk conservatism in availability tests holds even for waste material with relatively high dissolved organic carbon concentrations, which should induce increased release by ensuring the presence of an aquatic dispersive phase (73, 74). Sequential extraction schemes and tests that evaluate how leaching is affected by certain environmental parameters offer increased prediction capability. Sequential extraction schemes have not been used within waste policy support due to problems with selective extraction of the mineral phases (52). We shall return to sequential extraction schemes in the next section. Rather than applying a framework of operationally defined extractants, the current European waste characterization scheme relies on tests for pH and percolation. Leachate pH is a function of waste properties and local conditions, while percolation is influenced by local precipitation and the physical construction of the site. Sensitivity of the leaching rate to leachate pH is indicative of potential changes in metal mobility. The percolating volume increases with time, and percolation is therefore interpreted as the temporal element in the leaching process. The European Committee for Standardization has issued standard leaching tests for pH (79) and percolation (80). The relevant percolation volume for landfill situations in LCA would be infinite within the infinite time perspective of most LCA studies. However, observations show that as the cumulative percolation volume increases, the cumulative release of metal converges toward a value less than the total. The conversion value is equivalent to the potentially leachable fraction of metals. Given that near-infinite percolation is hard to investigate, the conversion is best illustrated through the change in leaching with pH, as illustrated in Figure 3. Release is controlled by the solubility of mineral phases in the waste and the availability of phases for attenuation and dispersion. Sequential Extraction. Metals in sediments and soils are associated with various phases, each mobilizable to various degrees and under certain conditions. Sequential extraction is the selective extraction of metal bound in target geochemical phases (81) and is thus often termed geochemical characterization. Sequential extraction is frequently used in risk characterization of soils and sediments to estimate the mobility of metals. Recent applications include solid waste fractions such as coal fly ash, waste incineration ash, sewage sludge, various waste fractions, and airborne dust (82, 83). With implications for many product life-cycles, residual

purpose of producing long-term leaching estimates, extraction results have to be interpreted according to the conditions that the material will experience. Most landfills undergo depletion of oxygen during a phase of degradation of organic constituents, followed by oxygenation at some point in time (26). The relevant basis for estimation of inventories in LCA therefore may include acid soluble carbonates as well as both oxidizable and reducible phases. Still, scenarios can be envisioned that never enter reducing conditions due to lack of degradable matter or if sedimentation is rapid. It is relevant to ask what the significance is of considering mobility and geoavailability in LCA. In order to exemplify the value of adding detail to long-term emission estimates based on knowledge of the mineral composition and characteristics of waste fractions we return to the diesel energy chain that we presented in the first section.

Synthesis

FIGURE 4. Interpretations met in the literature of metal mobility from sequential extraction. Notes refer to reference: a describes solids in general (82), b and c describe marine sediments (91, 92), d refers to marine drilling waste deposits (39), and e-h refer to soils (93–96) . metal, e.g., the least mobile inert metal fraction, is identified as the dominant phase for many metals in municipal solid waste (84). A list of the fractions identified by sequential extraction is summarized in Figure 4. Various schemes identify 3-9 different metal phases (59, 82, 85–87). Leachants for sequential extraction are designed for selectivity toward their target phase. Selective extraction of sulfides has shown to be particularly difficult as they to some extent will leach together with the other phases. Other sources of error include the sequence of extraction, incomplete dissolution of organics, organic coating on particulates, and redistribution during extraction (82, 85, 88). Investigation of the precision and repeatability of three extraction schemes showed variations within a factor of 2 for the extractable metal in each of the fractions (89). The most important issue in the experimental procedures was found to be the order in which the target phases were extracted, particularly significant for the organic and sulfide fraction. Ideally sulfides would be released during the extraction of the oxidizable fraction. Extraction of artificially prepared precipitates, however, shows that sulfides will be released both as part of the reducible fraction as well as the oxidizable fraction, and that the distribution between the two fractions depends on the material that is investigated and metal characteristics (90). Proper extraction of the exchangeable and oxidizable fraction is hampered by colloidal organics and in many cases an incomplete oxidation of total organics content combined with the difficulties in selectivity for the sulfide bound metals (82). Experimental conditions and sample preservation may also affect the outcome, particularly in the case of anoxic sediments. Due to the many influencing factors, sequential extraction is considered an operational rather than descriptive method (85, 88). Geoavailability and Mobility. As shown in Figure 4, several interpretations may be applied relating to the lability, mobility, or availability of metal phases estimated by sequential extraction. Some of the separation schemes apply a temporal perspective to mobilization. With reference to waste characterization, availability tests typically do not allow large changes to occur to the mineral matrix, such as changes in oxidization-reduction states. Geoavailable metal, interpreted here as long-term mobilizable metal, is generally considered the sum of all but the residual fraction. For the

Total, Geoavailable, and Highly Mobile Metal. We showed earlier in Figure 2 that metal releases from oil and gas drilling wastes contribute large shares to the diesel energy life-cycle ecotoxic potential. We shall use the case of drilling wastes to see how concepts of mobility and geoavailability may alter these conclusions. A large part of potentially hazardous metals in drilling wastes may be attributed to the weight agent in drilling chemicals. One class of weight agent consists of finely ground minerals added for density control. Most often the mineral is barite (BaSO4), but ilmenite (FeTiO3) is receiving increasing interest due to suspicions of Ba dissolution from barite. Many operators therefore prefer ilmenite when operating in sensitive areas. Both barite and ilmenite contain traces of hazardous metals, but while the mineral matrix of ilmenite is ecotoxically benign, barite is 50 wt-% Ba. In our assessment, the short-term mobile fraction is assumed represented by the leachable metal according to a pH-separated extraction scheme, cation-exchangeable and DTPA extractable fractions, and the water leachable fraction. All these are equal in concept, and assumed equivalent in our calculations, to the sum of exchangeable and carbonate fractions in sequential extraction schemes. In Figure 4 this is referred to as the highly mobile fraction. For geoavailable metal we separate a “high” and “low” estimate based on the sum of all fractions but the residual fraction, including or excluding the sulfide phase (oxidizable metal) respectively. Further details are reported in the Supporting Information. The release potential for total, geoavailable, and highly mobile metals bound in barite is summarized in Figure 5. Clearly, the estimated ecotoxic effect of deposition of barite in drilling waste is significantly affected by the level of risk conservatism applied to inventory estimation. Furthermore, even with good knowledge of the characteristics of barite mineral, large uncertainty remains in the leaching potentials due to both variability in mineral quality (i.e., total contents) and release potential (i.e., the highly mobile and geoavailable fractions). Implementation of a geoavailable fraction for metal release from marine deposition of drilling wastes in the diesel energy chain presented in Figure 2 leads to a decrease of a factor 50 for marine aquatic ecotoxicity compared to the original data, from constituting 50% of total marine aquatic ecotoxic potential to about 1%. In other words total marine aquatic ecotoxicity of diesel energy is halved, with possibly large consequences for life cycles that involve fossil energy. Implementation of geoavailable metal for the process of onshore treatment of drilling waste by land-farming does not affect the results to the same extent, with the contribution from this process decreasing from 33% of terrestrial ecotoxicity in the original inventory to 26% when geoavailable metal in barite is considered. VOL. 42, NO. 13, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 5. Uncertainty in total, geoavailable, and highly mobile metal in barite. Bars indicate lowest and highest values. The geometric mean is indicated for results based on more than 2 sources. For the geoavailable fraction: “high” assumes the residual fraction to be inert, while “low” considers both residual and oxidizable (sulfide phases) fractions inert.

FIGURE 6. Marine aquatic ecotoxic potential (MAETP; units of 1,4-dichlorobenzene, 1,4-DCB) of release to aquatic phase from 1 kg marine deposit of ilmenite and barite. Probability that barite performs better in terms of MAETP is indicated. Geoavailable “high” and “low” refer to releases including and excluding the sulfide phase (oxidizable phase) respectively; see Figure 5 and text for further explanation. Risk Perspectives. Total metal, the geoavailable fraction, and the highly mobile fraction may be viewed as the longterm release potential with different probabilities attached. The confidence in occurrence of release is as follows: P(release of total metal) < P(release of geoavailable fraction) < P(release of highly mobile fraction). Given that there are two options for the weight agent it is interesting to illustrate the relative ecotoxicity of barite versus that of ilmenite as a function of risk perspective. Again we rely on Huijbregts et al. for characterization factors (11). Marine deposition is assumed, and uncertainty in ecotoxic characterization factors is neglected. As outlined in Figure 6, the comparison of ilmenite versus barite is dominated by Ba bound in the mineral lattice of barite (BaSO4) when complete release is assumed. If the high estimate for geoavailability is applied (i.e., release of all phases except the residual), the conclusion is altered in favor of ilmenite. If the comparison is based on fractions that we are more confident of being released, the discerning power disappears and barite and ilmenite are considered equally ecotoxic in terms of metal release. In other words, the risk perspective that is applied has a large influence on the relative ecotoxic performance of the two alternatives. Take-Home Message. Metal ecotoxicity poses a challenge to LCA. Issues span from inventory estimation, through fate and exposure modeling, and ecotoxic effect assessment. In this review we focus on the emissions inventory estimation for long-term leaching processes, an issue of critical importance to many LCA studies. The traditional approach in LCA is to assume complete release of metal bound in solid waste, presuming that infinite time equals infinite weathering (22, 25, 41). Following this line of reasoning one may argue that within a truly infinite time-perspective, metals may be mobilized through weathering independent of human activity. There are several reasons for searching for better approaches. Studies of metal mobilization and immobilization show that there are processes which immobilize metals 4644

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from dissolved state and attenuate metals in solid state (97, 98). Maintaining that complete release occurs puts an inappropriate emphasis on long-term metal emissions in relation to instantaneous releases and degradable ecotoxic substances. This in effect installs a risk conservative bias into the inventory procedure for metals in solid wastes. Furthermore, complete release makes LCA biased against remediation of solids contaminated by metals. Current attempts at increasing the capability of LCA to assess long-term emissions all rely on use of a defined albeit arbitrary cutoff rule. Potentials for long-term release of metals in solid deposits are better discussed within a framework based on physical-chemical properties. Such approaches are well established within waste leaching tests and in the use of sequential extraction schemes. Within this framework we distinguish three levels of metal leaching potential: total metal, geoavailable metal, and highly mobile metal. In our opinion the concepts of mobility and geoavailability supply valuable information regarding the probability of future releases of metals in solid wastes and improve the quality of life-cycle inventory results.

Acknowledgments The International Research Institute of Stavanger (by Stig Westerlund) and Titania AS have kindly provided sequential extraction and leaching data for ilmenite and barite. StatoilHydro ASA is acknowledged for funding this work through a PhD grant.

Supporting Information Available Details regarding the diesel energy process, metal bound in barite and ilmenite. This information is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Heijungs, R.; Suh, S. The Computational Structure of Life Cycle Assessment; Kluwer Academic: Dordrecht, The Netherlands, 2002. (2) Baumann, H.; Tillman, A. M. The Hitch Hiker’s Guide to LCA An orientation in life cycle assessment methodology and application; Studentlitteratur: Lund, Sweden, 2004. (3) ISO 14040:2006. Environmental management - Life cycle assessment - Principles and Framework, 2nd ed.; International Organization for Standardization (ISO): Geneva, Switzerland, 2006. (4) Hauschild, M. Environmental impacts in a life-cycle perspective. Environ. Sci. Technol. 2005, 39 (4), 81A–88A. (5) Udo de Haes, H.; Finnveden, G.; Goedkoop, M.; Hauschild, M.; Hertwich, E.; Hofstetter, P.; Klo¨ppfer, W.; Krewitt, W.; Lindeijer, E.; Mueller-Wenk, R., et al. Life-Cycle Impact Assessment: Striving Towards Best Practice; SETAC Press: Pensacola, FL, 2002. (6) Scharnhorst, W.; Althaus, H.-J.; Classen, M.; Jolliet, O.; Hilty, L. M. The end of life treatment of second generation mobile phone networks: Strategies to reduce the environmental impact. Environ. Impact Assess. 2005, 25, 540–566. (7) Canals, L. M.; Burnip, G. M.; Cowell, S. J. Evaluation of the environmental impacts of apple production using life cycle assessment (LCA): case study in New Zealand. Agric. Ecosyst. Environ. 2006, 114, 226–238. (8) Moberg, Å.; Finnveden, G.; Johansson, J.; Lind, P. Life cycle assessment of energy from solid waste - Part 2: landfilling compared to other treatment methods. J. Cleaner Prod. 2005, 13, 231–240. (9) Gloria, T.; Russell, A. J.; Atherton, J.; Baker, S. R.; Cook, M. Ecological toxicity methods and metals: an examination of two case studies. Int. J. LCA. 2006, 11 (1), 26–33. (10) Huang, C.-C.; Ma, H.-W. A multidimensional environmental evaluation of packaging materials. Sci. Total Environ. 2004, 324, 161–172. (11) Huijbregts, M. A. J.; Thissen, U.; Guine`e, J. B.; Jager, T.; Kalf, D.; van de Meent, D.; van de Ragas, A. M. J.; Wegener Sleeswijk, A.; Reijnders, L. Priority assessment of toxic substances in life cycle assessment. Part 1: Calculation of toxicity potentials for 181 substances with the nested multi-media fate, exposure and effects model USES-LCA. Chemosphere 2000, 45, 659–669. (12) Payet, J.; Jolliet, O. Comparative assessment of the toxic impact of metals on aquatic ecosystems: the AMI method. In Life-Cycle Assessment of Metals: Issues and Research Directions, Dubreil, A., Ed.; SETAC Press: Pensacola, FL, 2002. (13) Goedkoop, M.; Spriensma, R. The Eco-Indicator 99. A Damage Oriented Method for Life Cycle Impact Assessment, 3rd ed.; Report 1999/36A; Ministerie van Volkshuisvesting, Ruimtelijke Ordening en Milieu: The Hague, The Netherlands, 2001. (14) Lollar, B. S. Environmental Geochemistry; Pergamon: Oxford, UK, 2005; Vol. 9. (15) Plumlee, G. S.; Logsdon, M. J. The Environmental Geochemistry of Mineral Deposits; Society of Economic Geologists: Littleton, CO, 1999; Vol. 6A. (16) Paquin, P. R.; Farley, K.; Santore, R. C.; Kavvadas, C. D.; Mooney, K. G.; Winfield, R. P.; Wu, K.-B.; DiToro, D. M. Metals in Aquatic Systems: a Review of Exposure, Bioaccumulation, and Toxicity Models; SETAC Press: Pensacola, FL, 2003. (17) Adams, W. J.; Chapman, P. M. Assessing the Hazard of Metals and Inorganic Metal Substances in Aquatic and Terrestrial Systems; CRC Press: Boca Raton, FL, 2007. (18) DiToro, D. M.; Kavvadas, C. D.; Mathew, R.; Paquin, P. R.; Winfield, R. P. The persistence and availability of metals in aquatic environments; International Council on Metals and the Environment (ICME): Ottawa, Canada, 2001. (19) Heijungs, R.; Koening, A. Improvement of LCA characterization factors and LCA practice for metals; TNO Environment, Energy and Process Innovation: Apeldoorn, The Netherlands, 2004. (20) Dubreil, A. Life-Cycle Assessment of Metals: Issues and Research Directions; SETAC Press: Pensacola, FL, 2002. (21) Doka, G.; Hischier, R. Waste treatment and assessment of longterm emissions. Int. J. LCA 2005, 10 (1), 77–84. (22) Hellweg, S. Time- and site-dependent life-cycle assessment of thermal waste treatment processes; dissertation.de - Verlag in Internet GmbH: Berlin, Germany, 2000. (23) Hellweg, S.; Hofstetter, T. B.; Hungerbu ¨hler, K. Time-dependent life-cycle assessment of slag landfills with the help of scenario analysis: the example of Cd and Cu. J. Cleaner Prod. 2005, 13, 301–320.

(24) Althaus, H.-J.; Classen, M. Life cycle inventories of metals and methodological aspects of inventorying material resources in ecoinvent. Int. J. LCA 2005, 10 (1), 43–49. (25) Finnveden, G. Methodological aspects of life cycle assessment of integrated solid waste management systems. Resour. Conserv. Recycl. 1999, 26 (3-4), 173–187. (26) Finnveden, G.; Albertsson, A.-C.; Berendson, J.; Eriksson, E.; Ho¨glund, L. O.; Karlsson, S.; Sundqvist, J.-O. Solid waste treatment within the framework of life-cycle assessment. J. Cleaner Prod. 1995, 3 (4), 189–199. (27) Ma, Y. B.; Uren, N. C. The effects of temperature, time and cycles of drying and rewetting on the extractability of zinc added to a calcerous soil. Geoderma 1997, 75, 89–97. (28) Hamon, R. E.; McLaughlin, M. J.; Naidu, R.; Correll, R. Longterm changes in Cadmium bioavailability in soil. Environ. Sci. Technol. 1998, 32 (23), 3699–3702. (29) Ma, Y.; Lombi, E.; Nolan, A. L.; McLaughlin, M. J. Short-term natural attenuation of copper in soils: effects of time, temperature, and soil characteristics. Environ. Toxicol. Chem. 2006, 25 (3), 652–658. (30) Young, S. D.; Crout, N.; Hutchinson, J.; Tye, A.; Nakhone, L. Techniques for measuring attenuation: isotopic dilution methods. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007. (31) Chapman, P. M. Inorganic metals and metalloids ERA: recent advances and implications for LCIA. In Life-Cycle Assessment of Metals: Issues and Research Directions; Dubreuil, A., Ed.; Society of Environmental Toxicology and Chemistry (SETAC): Pensacola, FL, 2005. (32) Eckelman, M. J.; Graedel, T. E. Silver emissions and their environmental impacts: a multilevel assessment. Environ. Sci. Technol. 2007, 41 (17), 6283–6289. (33) Hertwich, E. G.; Hammitt, J. K. A decision-analytic framework for impact assessment. Part 1: LCA and decision analysis. Int. J. LCA 2001, 6 (1), 5–12. (34) Green, A. S.; Chapman, P. M.; Allen, H. E.; Campbell, P. G. C.; Cardwell, R. D.; de Schamphelaere, K.; Delbeke, K. M.; Mount, R. D.; Stubblefield, W. A. Aquatic toxicity for hazard identification of metals and inorganic metal substances. In Assessing the Hazard of Metals and Inorganic Metal Substances in Aquatic and Terrestrial Systems; Adams, W. J., Chapman, P. M., Eds.; CRC Press: Boca Raton, FL, 2007. (35) Schlekat, C. E.; McGeer, J. C.; Blust, R.; Borgmann, U.; Brix, K. L.; Bury, N.; Couillard, Y.; Dwyer, R. L.; Louoma, S. N.; Robertson, S., et al. Bioaccumulation: hazard identification of metals and inorganic metal substances. In Assessing the Hazard of Metals and Inorganic Substances in Aquatic and Terrestrial Systems; Adams, W. J., Chapman, P. M., Eds.; CRC Press: Boca Raton, FL, 2007. (36) Smith, K. S.; Huyck, H. L. O. An overview of the abundance, relative mobility, bioavailability, and human toxicity of metals. In The Environmental Geochemistry of Mineral Deposits. Part A: Processes, Techniques, and Health Issues; Plumlee, G. S., Logsdon, M. J., Eds.; Society of Economic Geologists: Littleton, CO, 1999; Vol. 6A. (37) Newman, M. C.; Unger, M. A. Fundamentals of Ecotoxicology, 2nd ed.; Lewis Publishers: London, UK, 2002. (38) Frischknecht, R.; Jungbluth, N.; Althaus, H.-J.; Doka, G.; Dones, R.; Hellweg, S.; Hischier, R.; Nemecek, T.; Rebizer, G.; Spielmann, M. Ecoinvent v1.1; Swiss Centre for Life Cycle Inventories: Uster, Switzerland, 2004. (39) Neff, J. M. Composition, Environmental Fates, and Biological Effect of Water Based Drilling Muds and Cuttings Discharged to the Marine Environment: A Synthesis and Annotated Bibliography; Prepared for Petroleum Environmental Research Forum (PERF) and American Petroleum Institute (API); Batelle: Duxbury MA, 2005. (40) Finnveden, G.; Nielsen, P. H. Long-term emissions from landfills should not be disregarded. Int. J. LCA 1999, 4 (3), 125–126. (41) Udo de Haes, H. A.; Jolliet, O.; Finnveden, G.; Hauschild, M.; Krewitt, W.; Mu ¨ ller-Wenk, R. Best available practice regarding impact categories and category indicators for life cycle impact assessment, Part 1. Int. J. LCA 1999, 4 (2), 66–74. (42) DiToro, D. M.; Allen, H. E.; Bergman, H. L.; Meyer, J. S.; Paquin, P. R. Biotic ligand model of the acure toxicity of metals. 1: Technical basis. Environ. Toxicol. Chem. 2001, 20 (10), 2383–2396. (43) Rieuwerts, J.; Thornton, I.; Farago, M.; Ashmore, M. Quantifying the influence of soil properties on the solubility of metals by predictive modelling of secondary data. Chem. Speciation Bioavailability 1998, 10 (3), 83–94. VOL. 42, NO. 13, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4645

(44) Rieuwerts, J.; Thornton, I.; Farago, M.; Ashmore, M. Factors influencing metal bioavailability in soils: preliminary investigations for the development of a critical loads approach for metals. Chem. Speciation Bioavailability 1998, 10 (2), 61–75. (45) Baker, S.; Herrchen, M.; Hund-Rinke, K.; Klein, W.; Ko¨del, W.; Peijneburg, W.; Rensing, C. Underlying issues including approaches and information needs in risk assessment. Ecotoxicol. Environ. Saf. 2003, 56, 6–19. (46) Oorts, K.; Ghesquiere, U.; Swinnen, K.; Smolders, E. Soil properties affecting the toxicity of CuCl2 and NiCl2 for soil microbial processes in freshly spiked soils. Environ. Toxicol. Chem. 2006, 25 (3), 836–844. (47) Unsworth, E. R.; Warnken, K. W.; Zhang, H.; Davison, W.; Black, F.; Buffle, J.; Cao, J.; Cleven, R.; Galceran, J.; Gunkel, P. Model predictions of metal speciation in freshwaters compared to measurements by in situ techniques. Environ. Sci. Technol. 2006, 40, 1942–1949. (48) Astrup, T.; Dijkstra, J. J.; Comans, R. N. J.; van der Sloot, H. A.; Christensen, T. H. Geochemical model of leaching from MSWI air-pollution-control residues. Environ. Sci. Technol. 2006, 40, 3551–3557. (49) Reimann, C.; Garret, R. G. Geochemical background - concept and reality. Sci. Total Environ. 2005, 350, 12–27. (50) Hall, J. R.; Ashmore, M.; Fawehinmi, J.; Jordan, C.; Lofts, S. Developing a critical load approach for national risk assessments of atmospheric metal deposition. Environ. Toxicol. Chem. 2006, 25 (3), 883–890. (51) Olsen, S. I.; Christensen, F. M.; Hauschild, M.; Pedersen, F.; Larsen, H. F.; Tørsløv, J. Life cycle impact assessment and risk assessment of chemicals - a methodological comparison. Environ. Impact Assess. 2001, 21 (4), 385–404. (52) Kosson, D. S.; van der Sloot, H. A.; Sanchez, F.; Garrabrants, A. C. An integrated framework for evaluating leaching in waste management and utilization of secondary materials. Environ. Eng. Sci. 2002, 19 (3), 159–203. (53) Lloyd, S. M.; Ries, R. Characterizing, propagating, and analyzing uncertainty in life-cycle assessment. J. Ind. Ecol. 2007, 11 (1), 161–179. (54) van der Sloot, H. A.; Heasman, L.; Quevauviller, P. Harmonization of Leaching/Extraction Tests; Elsevier Science: Amsterdam, The Netherlands, 1997. (55) Hellweg, S.; Hofstetter, T. B.; Hungerbu ¨ hler, K. Modeling waste incineration for life-cycle inventory analysis in Switzerland. Environ. Model. Assess. 2001, 6, 219–235. (56) Goldberg, E. D. Minor elements in sea water. In Chemical Oceanography; Riley, J. P., Skirrow, G., Eds.; Academic Press: London, UK, 1965. (57) Whitfield, M. The world ocean: mechanism or machination. Interdiscipl. Sci. Rev. 1981, 6 (1), 12–35. (58) Menzies, N. Redox processes and attenuation of metal availability in soil. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007. (59) Tack, F. M.; Verloo, M. G. Chemical speciation and fractionation in soil and sediment heavy metal analysis: a review. Int. J. Environ. Anal. Chem. 1995, 59, 225–238. (60) Byrne, R. H., Speciation in seawater. In Chemical Speciation in the Environment, 2nd ed.; Ure, A. M., Davidson, C. M., Eds.; Blackwell Science Ltd: Oxford, UK, 2002. (61) Elderfield, H. The oceanic chemistry of the rare earth elements. Philos. Trans. R. Soc. London, A 1988, A325, 105–126. (62) Impelliteri, C.; Allen, H. E., Effects of humic substances on attenuation of metals: bioavailability and mobility in soil. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007. (63) McBride, M. B. Attenuation of metal toxicity in soils by biological processes. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007. (64) Mench, M.; Vangronsveld, J.; Lepp, N.; Ruttens, A.; Bleeker, P.; Geebelen, W. Use of soil amendments to attenuate trace element exposure: sustainability, side effects, and failures. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007. (65) Bayer, P.; Finkel, M. Life cycle assessment of active and passive groundwater remediation technologies. J. Contam. Hydrol. 2006, 83, 171–199. (66) Volkwein, S.; Hurtig, H.-W.; Klo¨pffer, W. Life cycle assessment of contaminated sites remediation. Int. J. LCA 1999, 4 (5), 263– 274. 4646

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 42, NO. 13, 2008

(67) Blanc, A.; Me´tivier-Pignon, H.; Gourdon, R.; Rousseaux, P. Life cycle assessment as a tool for controlling the development of technical activities: application to the remediation of a site contaminated by sulfur. Adv. Environ. Res. 2004, 8, 613–627. (68) Page, C. A.; Diamond, M. L.; Campbell, M.; McKenna, S. Lifecycle framework for assessment of site remediation options: case study. Environ. Toxicol. Chem. 1999, 18 (4), 801–810. (69) Wrisberg, N.; Udo de Haes, H. A.; Triebswetter, U.; Eder, P.; Clift, R. Analytic Tools for Environmental Design and Management in a Systems Perspective: The Combined Use of Analytical Tools; Kluwer Academic Publishers: Dordrecht, The Netherlands, 2002. (70) Nielsen, P. H.; Hauschild, M. Product specific emissions from municipal solid waste landfills. Part 1: Landfill model. Int. J. LCA 1998, 3 (3), 158–168. (71) Hofstetter, P. Perspectives in Life Cycle Impact Assessment. A Structured Approach to Combine Models of the Technosphere, Ecosphere and Valuesphere; Kluwer Academic Publishers: Norwell, MA, 1998. (72) Hellweg, S.; Hofstetter, T. B.; Hungerbu ¨ hler, K. Discounting and the environment. Int. J. LCA 2003, 8 (1), 8–18. (73) van der Sloot, H. A.; Comans, R. N. J.; Meeussen, J. C. L. J. D. J. Leaching Methods for Soil, Sludge and Treated Biowaste. Final Report; Desk study no. 23; Project HORIZONTAL (www.ecn.nl/ horizontal), online research network aimed to develop horizontal and harmonised European standards in the field of sludge, soil, and treated biowaste; Maintained by Energieonderzoek Centrum Nederland (ECN): Petten, The Netherlands, 2004. (74) Grathwohl, P.; Halm, D. E. Guideline for Groundwater Risk Assessment at Contaminated Sites (GRACOS); Report of the GRACOS research project of the European Commission, implemented under the Fifth Framework Programme; Print: Campus Druck: Tu ¨ bingen, Germany, 2003. (75) NEN. NEN 7341: Leaching characteristics of solid earthy and stony building and waste materials - Leaching tests - Determination of the availability of inorganic components for leaching; Netherlands Normalisation Institute: Delft, The Netherlands, 1995. (76) Nordtest . NT Envir 003: Solid waste - Granular inorganic material - Availability test; Nordtest: Espoo, Finland, 1995. (77) Skeaff, J. M.; McGeer, J. C.; King, P.; King, M.; Nadeau, J. C. On the aquatic hazard classification of metals and alloys. In LifeCycle Assessment of Metals: Issues and Research Directions; Dubreuil, A., Ed.; Society of Environmental Toxicology and Chemistry (SETAC): Pensacola, FL, 2005. (78) OECD. Guidance document on transformation/dissolution of metals and metal compounds in aqueous media. OECD series on testing and assessment, No. 29; Organization for Economic Co-operation and Development; Environment, Health and Safety Division: Paris, France, 2001. (79) CEN. CEN/TS 14429: Characterization of waste - Leaching behaviour tests - Influence of pH on leaching with initial acid/ base addition; European Committee for Standardization: Brussels, Belgium, 2005. (80) CEN. CEN/TS 14405: Characterization of waste - Leaching behaviour tests - Up-flow percolation test (under specified conditions); European Committee for Standardization: Brussels, Belgium, 2004. (81) Tessier, A.; Campbell, P. C. G.; Bisson, M. Sequential extraction procedure for the speciation of particulate trace metal. Anal. Chem. 1979, 51 (7), 844–851. (82) Filgueiras, A. V.; Lavilla, I.; Bendicho, C. Chemical sequential extraction for metal partitioning in environmental solid samples. J. Environ. Monit. 2002, 4, 823–857. (83) Ure, A. M.; Davidson, C. M. Chemical speciation in soils and related materials by selective chemical extraction. In Chemical Speciation in the Environment, 2nd ed.; Ure, A. M., Davidson, C. M., Eds.; Blackwell Science Ltd: Oxford, UK, 2002. (84) Prudent, P.; Domeziel, M.; Massiani, C. Chemical sequential extraction as decision-making tool: application to municipal solid waste and its individual constituents. Sci. Total Environ. 1996, 178, 55–61. (85) Tessier, A.; Campbell, P. C. G. Partitioning of trace metals in sediments: relationships with bioavailability. Hydrobiologia 1987, 149, 43–52. (86) Ure, A. M. Single extraction schemes for soil analysis and related applications. Sci. Total Environ. 1996, 178, 3–10. (87) Sahuquillo, A.; Rigol, A.; Rauret, G. Overview of the use of eaching/extraction testst for risk assessment of trace metals in contaminated soils and sediments. Trends Anal. Chem. 2003, 20 (10), 2397–2402.

(88) Pickering, W. F. General strategies for speciation. In Chemical Speciation in the Environment, 2nd ed.; Ure, A. M., Davidson, C. M., Eds.; Blackwell Science Ltd: Oxford, UK, 2002. (89) Ure, A. M.; Quevauviller, P.; Muntau, H.; Griepink, B. Speciation of the heavy metals in soils and sediments - an account of the improvement and harmonization of extraction techniques undertaken under the auspices of the BCR of the commission of the European communities. Int. J. Environ. Anal. Chem. 1993, 24, 1573–1578. (90) Batley, G. E. Heavy metal speciation in waters, sediments and biota from Lake Macquarie, New South Wales. Aust. J. Mar. Freshw. Res. 1987, 38, 591–606. (91) Sa´enz, V.; Blasco, J.; Go´mez-Parra, A. Speciation of heavy metals in recent sediments of three coastal ecosystems in the Gulf of Ca´diz, Southwest Iberian Peninsula. Environ. Toxicol. Chem. 2003, 22 (12), 2833–2839. (92) Riba, I.; DelValls, T. A.; Forja, J. M.; Go´mez-Parra, A. Influence of the Aznalco´llar mining spill on the vertical distribution of heavy metals in sediments from the Guadalquivir estuary (SW Spain). Mar. Pollut. Bull. 2002, 44, 39–47. (93) John, D. A.; Leventhal, J. S. Bioavailability of metals. In Preliminary Compilation of Descriptive Geoenvironmental Min-

(94)

(95) (96)

(97) (98)

eral Deposit Models; du Bray, E. A., Ed.; U.S. Department of the Interior and US Geological Survey: Denver, CO, 1995. Manz, M.; Weissflog, L.; Ku ¨ hne, R.; Schu ¨u ¨ rmann, G. Ecotoxicological hazard and risk assessment of heavy metal contents in agricultural soils of central Germany. Ecotoxicol. Environ. Saf. 1999, 42, 191–201. Almås, Å.; Singh, B. R.; Salbu, B. Mobility of Cadmium-109 and Zinc-65 in soil influenced by equilibrium time, temperature, and organic matter. J. Environ. Qual. 1999, 28, 1742–1750. Kennedy, V. H.; Sanchez, A. L.; Oughton, D. H.; Rowland, A. P. Use of single and sequential chemical extractants to assess radionuclide and heavy metal availability from soils for root uptake. Analyst 1997, 122, 89R–100R. Yong, R. N.; Mulligan, C. N. Natural Attenuation of Contaminants in Soil; CRC Press: Boca Raton, FL, 2004. Singh, B. R. Natural attenuation of trace element availability assessed by chemical extraction. In Natural Attenuation of Trace Element Availability in Soils; Hamon, R. E., McLaughlin, M. J., Lombi, E., Eds.; CRC Press: Boca Raton, FL, 2007.

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