Open access peer-reviewed chapter

Heavy Metal Speciation, and the Evaluation and Remediation of Polluted Mine Wastes and Soils

Written By

Arturo Aguirre Gómez and Margarita Eugenia Gutiérrez Ruiz

Submitted: 24 December 2022 Reviewed: 07 February 2023 Published: 06 March 2023

DOI: 10.5772/intechopen.110412

From the Edited Volume

Heavy Metals - Recent Advances

Edited by Basim A. Almayyahi

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Abstract

The chapter exposes how a sound methodology can be instrumented to both, biogeochemically speciate heavy metal (HM) polluted mine wastes and soils, and to develop solid strategies to agriculturally stabilize and remediate HM-polluted terrestrial environments. Using single- and sequential extraction procedures, polluted environments can be chemically speciated to successfully remediate impacted sites. Once metal(loid) toxic levels are determined, common amendments (compost, P-fertilizers, lime, gypsum) can be added to abate HM levels, and to re-sustain vegetation, based on bioassay results of HM-sensitive plants. The approach addresses first: a) a discussion of concepts and relevant chemistry that apply to study mine tailing materials and soils, via single or multiple HM-fractionation schemes; b) characterizing chemically mine tailings and soils, in terms of the metal(loid)-sorption-complexing affinities, and c) creating a “fertile environment” by agriculturally reconditioning the HM-polluted acidic mine waste to allow the vegetation regrowth, based on bioassay test performance. Results of two successful cases of study are included; one showing the use of single extraction procedures to evaluate phytoavailable/toxic HM levels to agriculturally remediate polluted sites, and another showing the role of sequential extraction procedures to discriminate heavy metal(loid)s of a spill from other metal deposits of the same ore.

Keywords

  • soil pollution
  • metal mine tailings
  • chemical speciation
  • heavy metalloid bioremediation
  • heavy metal bioassays

1. Introduction

Mining extraction of heavy metals from sulfidic materials produces considerable levels of potential acidity which eventually, if not prevented and neutralized generate the so-called metalliferous acidic mine drainage, resulting in a potential mobilization of soluble heavy metals [1, 2]. The quantity of acid-forming minerals found in many mines of Central Mexico around the neo-volcanic axis [3, 4] and in Northern Mexico contains dominantly pyrite (FeS2), galena (PbS), sphalerite (ZnS), pyrrhotites (Fe1-xS), chalcopyrite (CuFeS2), arsenopyrite (AsFeS), bornite (Cu5FeS4), and many other metallic sulfosalts [3]. After oxidation, these minerals generate the H+-producing redox and hydrolytic processes of the components (e.g., S, Fe, Mn, Zn, Cu, Pb, etc.) left behind in the mine tailings. Mine wastes, polluted- and pristine-soils must then be studied and chemically speciated [5], usually by applying simple or multiple-sequential extraction procedures [5, 6, 7, 8], this to fractionate the HM species according to their expected chemical interaction with the various solid phase compartments present in soils or mining wastes. Based on that, HMs can be grouped into one or several of the following categories: a) water-soluble (free metal ions, M2+, inorganic and organic metal complexes, ML, whether labile or not, etc.); b) exchangeable (non-specifically adsorbed); c) ligand extractable (bioavailable); d) acid-extractable (carbonate-precipitated); e) organically and sulfide bound (oxidizing fraction); f) chemisorbed on Fe-, and Mn oxides (reducible fraction) and g) lattice-retained (occluded or residual). Metal speciation and fractionation to assess any remediation strategy must then rely on finding the right methodology to evaluate which of the extraction procedures may serve as a metal available/toxic fraction indicator [5, 6, 7, 8]. This will eventually assure that a metal-sensitive plant can grow once any other undesirable physicochemical characteristics are resolved, such as low pH values, low nutrient status, and high EC). The proposal must then include the addition of ameliorating materials (lime, phosphates, compost, gypsum) that both, help to abate the HM-phytoavailable levels, while serving to neutralize the acidified metal-polluted site, and to mend the growing media, so that plant regrowth will not be impeded. Once these two aspects are guaranteed, results must, all in all, meet the requirements of national or international standards and norms and pass the chosen specific bioassay so that the site can be considered agrostabilized [9, 10, 11]. Other necessary physical and chemical characterization procedures shall include the total metal content, [MT], original minerals identification, electrical conductivity, pH, HM-sorption-complexing affinities, etc. [12, 13]. The chapter addresses in Section 2, the relevant chemistry to be considered when researching HMs in mine tailings and polluted soils to chemically characterize wastes and pristine soils. The most commonly chemical speciation and fractionation terms and protocols used are discussed in terms of validating which one best adapts to the specific purpose. In Sections 3 and 4, two cases of study are presented: A) One including an example of how a single extraction procedure was used to study-remediate a Cd-, Cu-, Pb- and Zn- polluted mine waste from “Mina La Negra”, at Zimapan, Hidalgo State, in Central Mexico, following a successful application of a methodology to: a) sufficiently abate the levels of phytoavailable-toxic HM-contents of the mine waste by adding lime, gypsum, phosphates or compost; b) ameliorate the waste material to create a “fertile” environment and; c) test a) and b) by applying and passing a well-established bioassay using an HM-sensitive plant, this to assure a successful regrow of vegetation; and B) Another, including a successful application of a sequential extraction protocol to evaluate the effect a spill of a Cu-mine into the Sonora River basin in Northern Sonora State, Mexico, on the bio-accessible levels of metal(loid)s such as Fe, Al, Cu, Mn, and minor amounts of As, Cd, Zn, Pb, and Cr.

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2. Chemical speciation and extraction procedures

It is well known that the chemical speciation of heavy metal(loid)s depends not on their total concentrations, but more on the form in which they are found in the environment [5]. Chemical forms depend not only on the reactions that control HM solubility [14, 15, 16], and thus on their availability [17, 18, 19] and toxicity to plants [20], but, on their mobility and distribution in the environment [21], which in turn are dominated by their physicochemical interactions (complexation, redox-chemistry, and sorption) [14] with the different solid phase compartments (minerals lattices, oxides, organics, carbonates). Based on this, it is important first to discuss the relevant chemistry of these processes, and secondly, to which extent these will give a good indication of the availability/toxic levels to plants or other living organisms. Because many international standards [22] and national norms [23, 24, 25, 26, 27, 28] base their threshold-legal values on certain metal-extractabilities, and in few cases, on the performances of certain bioassays and biological tests [29, 30, 31, 32, 33, 34], a good selection of extracting procedures and pertinent bioassays must be correlated to assure both, the successful remediation strategy followed, and the health of the terrain to sustain vegetation regrowth. Although several research papers and reviews have been published in the past decades regarding the importance of chemical speciation studies and sequential extraction protocols [5, 6, 7, 8] to evaluate sediments, soils, biosolids, etc. [5] there is still a strong need to establish a sound methodology, in terms of their applicability, to examine the chemistry of HM-polluted mine wastes. To do so, it is important first then, to present a focused review of the relevant chemical aspects of the interactions implicated among the diverse metal(loid)s present, and the various solid phases found on these wastes.

2.1 Chemical speciation; definitions and concepts

Although the term “heavy metal” could be imprecise and sometimes misleading, as stated by IUPAC [35], over the past four decades, the term “heavy metal” continues to be widely used and applied to a group name of metal(loid)s and that are associated with pollution and/or potential (eco)toxicity. Besides, legal regulations often refer to the term “heavy metal” not only not specifying which metals are included, but under which chemical basis it is assumed that HM and their compounds have (eco)toxic effects, or pose analogous physicochemical, biological, and toxicological properties. Thus, any new name, definition, or classification of metals would be better based on the chemical properties of metals, and such a categorization should reflect our understanding of the chemical basis of reactivity and toxicity, so that their toxic effects can be predicted. More appropriate terms seem to be sound if based on the relevant chemical behavior of the metal, such as those considered by the hard and soft acid and base Pearson’s-Lewis theory [36, 37, 38, 39, 40, 41], or the “s”, “p”, “d”, and “f” character of metals, according to the orbitals used in the HM-bonding, or on their relevant periodic properties involved in the interaction (electronegativity, ionic potential, ionic radius, metal hydrolysis)[14] that governs their ultimate fate. Other names have been suggested [5, 6, 7, 8, 42] presumably being more appropriate, such as potentially toxic elements, PTEs [42]. Yet, although the name heavy metal persists and continues to be used in literature, policies, and regulations, and we will utilize it, in this chapter, in correspondence with the title of the book, and following the general use and acceptance of many researchers in the literature, reference will be made to the chemical aspects of metal reactivity.

2.1.1 Chemical, functional, and operational speciation

Several attempts have been made to clarify unambiguously the term chemical speciation (of heavy metals) in environmental and agricultural sciences [5, 7, 35, 42], especially when used to characterize the relevant chemistry of a specific element that can be toxic or nutrimental to living organisms. However, when a metal is distributed in the environment, several ways are used to describe its behavior in the environment, or its chemical form and activity, among other properties. In the following sections, we will focus not only on the metal chemical forms by itself, in terms of the phases where the metal is distributed (aqueous or solid), but unambiguously on the nature of the specific bonding involved (van der Waals forces, ionic or covalent bonds, inner or outer sphere complexes, etc.), and on the type of compartment that the specific HM occupies, according to its interaction with the solid phase, such as water-soluble, exchangeable (non-specifically adsorbed), chemisorbed (specifically adsorbed), chelate extractable (available), acid- extractable (carbonate-precipitated), organically+sulfide bound (oxidizable), Fe/Mn oxides (reducible), occluded-residual (lattice-retained), or even more, on the ability to be taken up, ingested, bioaccumulated, etc., by living organisms. In a document on chemical speciation terminology published by IUPAC [43], it is recommended the term ‘chemical species’ for describing the form of an element defined as isotopic composition, electronic or oxidation state, and/or complex or molecular structure. This definition has been considered as inappropriate for most studies on solid materials such as soils, sediments, and other materials [5, 42] and transfers the selective extraction procedures relevant to such solid phases to the category of fractionation methodologies. Broader definitions to include soils and mine wastes, for instance, are defined in terms of the so-called functional or operational speciation. In this respect, the term may be better defined to identify, describe, and quantify the amounts of the species forms and phases present in each material [7, 42, 43]. Thus, a general consensus points toward adopting a definition that includes both, forms or phases, so that speciation seems to be better defined as: a) referring to a specific chemical compound or oxidation state in environmental samples, even though this type of speciation in which the precise chemical form of an element is measured is the most difficult to achieve, since very sensitive and selective analytical techniques are required; and; b) functional speciation, for specific usages as for terms like phytoavailable (plant-available species) or bio-accessible (for animals or humans), etc., or c) operational speciation, defined in terms of the extraction procedure utilized to refer to the physical or chemical fraction characterized (water-soluble, acid-extractable, etc.) in soils, or mine wastes. Physical procedures to divide samples by particle size, or fractions separated by filtration, centrifugation, or dialysis, etc., are also considered an operational speciation [7, 42, 44], as the distinction between soluble and insoluble species is based on the ability to pass a sample through a 0.45 μm filter. In fact, many operational procedures are often used to fraction metals based on particle size: dissolved (<1 kDa), colloidal (1 kDa-0.45 μm), and particulate (> 0.45 μm).

2.1.2 The heavy metal-solid phase interaction: Extraction protocols and fractions

It has been long recognized that knowledge of both, the elemental composition of the solid materials (sediments, soils, biosolids, mine tailings, etc.), and the total concentration of HMs present in the environment, are of little use in assessing the availability and toxicity to biota. For these reasons, chemical tests have relied more on measurement of extractable or “labile” fractions of these potentially toxic elements. Such tests, however, have provided little basis to relate HM-extractability in mine wastes, to the chemical forms that can be toxic to organisms and deleterious to the environment. To evaluate and plan a remediation strategy for an HM-polluted site, a fundamental understanding of the processes that control heavy metal solubility and availability to biota is needed. Thus, the relevant physicochemical aspects of the interactions between HM and the solid phases present in the mine waste deposits must be reviewed to successfully correlate their extractability with plant tolerance, so that both, the requirements imposed by standards and norms are fulfilled, and the site can be agriculturally stabilized to allow revegetation. Chemical interactions among HMs and solid phases in unaltered minerals of soils, and in altered mine wastes left behind after ore exploitation, pose and/or create very diverse physical and chemical conditions that influence speciation, such as 1) ionic strength-electrical conductivity of media, 2) presence of dissolved organic matter and complexing ligands, 3) pH and potential acid-forming equilibria, 4) redox potential, Eh-pe-values, 5) hard/soft-acid/base character of the metals (charge, ionic radii, ionic potential, metal hydrolysis, etc.), and 6) reaction kinetics. The combined action of these factors, plus those of the edaphogenetic characteristics of the original soil, the biogeochemical processes occurring, and climate will favor the formation of different metal species, resulting in an innocuous/toxic HM species with a higher or lower bioavailability [45]. Thus, analyzing metal concentrations of aqueous or solid phases is not sufficient to determine its toxicity as biological effects [46], so that, chemical knowledge might provide a more effective diagnosis of environmental conditions [47]. Several biological factors that may influence the bioavailability of metals, include the route of exposure, the mechanism of sequestration and transport of metals by organic ligands, and the exposed organism [48]. According to Rainbow and Luoma [49], in metal ecotoxicology, the term bioavailability, corresponds, both, to the metal that is available for capture by a living organism and is integrated into the metabolic processes, and the fraction of the concentration of the metal that is absorbed and/or adsorbed by the organism. The assimilated fraction may then interact with receptors and physiological sites, causing toxic effects [49]. In the following sections a brief review of the relevant chemistry that will otherwise tilt the balance toward a given single or sequential extraction procedures to speciate HM in mine wastes and soils, will be discussed. Based on Tessier et al. [6], and other studies [5, 42, 50] and protocols [7, 8], the most common soil- and mine waste species and phases to be single or sequentially extracted may include: a) the soil- and mine waste-water-soluble fraction; b) the exchangeable/non-specifically adsorbed species; c) the acid-extractable (carbonate+specifically chemisorbed species) phase; d) the (oxidizable) organically complexed+sulfide metal species; e) the (reducible) hydrous Fe/Mn oxides fraction; and f) the residual-occluded and strong acid-extractable species.

2.2 The water-soluble fraction and solution speciation

In the soil, sediment, and mine waste solutions, the chemical speciation of heavy metals (HM) must consider the solvation process and complexes formed with organic and inorganic ligands [51]. In solution, ligands can form inner- or outer sphere complexes with metal cations [52]. Information on solution speciation is required for predicting bioavailability because the free metal ion, M2+, is the most toxic species for biota and the most reactive one that interacts with the solid phases [53]. Depending on the metal, the free ion may be the dominant species (Cd2+(aq), Zn2+(aq), etc.) or may account for only a minor fraction of the total metal in solution [54], due to the formation of stable metal complexes forms (CuL, PbL, etc.). Metals in soil solution may be present as M2+(aq), or as kinetically labile/non-labile metal- complexes, with inorganic (Cl, SO42−, CO32−, etc.), or organic ligands, or associated with mineral colloids [52, 53]. Thus, the analytical evaluation of the free metal activity, (M2+)-value, is an important step in the process of chemical speciation of pollutant metals in aqueous solutions [38, 55, 56, 57, 58]. However, in the past, the experimental determination of (M2+) was restricted due to several limitations regarding sensitivity and selectivity of the applied methodologies. Traditionally, the strict evaluation of (M2+) was only possible using solid-state metal ion-selective electrodes (M-ISE), but some other analytical techniques have also been used to estimate ‘free ion’ concentrations of metals in solution (Donnan dialysis, resin exchange methods, and chromatographic techniques [54, 59], voltammetry (e.g., Anodic Stripping Voltammetry, ASV [56], or Cathodic-Stripping Voltammetry)). However, HM ion fraction determinations, for instance, might become unreliable if concentrations in solution are below the detection limits (e.g., Donnan dialysis, resin exchange methods) necessary for the use of very sensitive techniques such as ASV [55, 56]. Measurements with ion-selective electrodes (ISEs) in natural samples may be affected by the fouling of the electrode by organic matter [60]. The use of ISEs under large Cl concentrations are also not advisable, and other interferences may occur, resulting in an overestimation of the free metal ion concentrations [61]. Although during the last decades, much progress has been made in reducing the detection limits of ISEs [62]. Most studies to measure trace metal speciation in soil solutions have been on Cu2+, for which detection limits of 10−13 M or even smaller have been reported [63]. Voltametric techniques, specifically ASV, have been successfully used to determine not only the concentrations of labile metal species from a current measured in solution as a metal is dissolved or released from a Hg-electrode, but the (M2+)-value too. Although analytical problems have been claimed regarding overlapping of stripping peaks, adsorption of surface-active organic compounds on the Hg-surface inhibiting the metal deposition, or formation of insoluble intermetallic compounds that affect peak size and position, other studies showed that these inconveniences can be overcome, so that free Cd2+, Cu2+, Pb2+, and Zn2+ activities [55, 56] can be measured in natural polluted soil solution samples.

2.2.1 The free metal activity measurement

Traditionally, the strict evaluation of (M2+)-values was only possible using solid-state metal ion-selective electrodes (M-ISE), but except perhaps for Cu2+, no M-ISE has adequate sensitivity and specificity for evaluating trace (free) metals in solutions where many metals coexist. It is well known that the M-ISE for Cd2+ and Pb2+, for instance, respond similarly to both metals; hence, Cd and Pb interfere with each other [64, 65]. Aguirre et al. [55, 56] developed a robust method to determine (M2+)-values of Cd, Cu, Pb, and Zn by ASV. The method was tested using metal-buffer solutions to control (M2+), by complexing metals with weak, medium, and strong ligands, and varying pH, total metal aqueous concentration (10−6–10−7 M), metal–ligand ratios of 1:20 M, and 0.010 M acetate medium. For the studied metals, Cd, Cu, Zn, and Pb, the agreement was found among theory, experimental ASV measurements, and (M2+), predicted by using a speciation chemical equilibria program and stability constants reported in the literature. Good agreement was found between the theory and calculated (M2+), and between experimental ASV results and calculated (M2+). Free metal activities in the order of pCd ≤ 12, pCu ≤ 18, pPb ≤ 10, and pZn ≤ 9 were measurable under the established experimental conditions. Results (not shown) also revealed good agreement between Cu-ISE and Cu-ASV when measuring the free Cu2+ activity in aqueous extracts of four soils. Values of soil-(Cu2+) measured were in the order of 10−5 to 10−9. The calibration curves for each metal were prepared according to speciation calculations of the metal–ligand–pH equilibrium systems (M–L–pH) in 0.1 M acetate as an indicator ligand, since fulvic acids in soils contain appreciable amounts of carboxylic compounds of low molecular weights, such as acetates, oxalates, and citrates, among others. Synthetic solutions of final concentrations of total metal [MT] = 1 × 10−5 M, total ligand [LT] =2–4 × 10−4 M, and pH values in the 4.0–7.5 range were tested. The calibration curves (ΔEp)c versus log(M2+) were generated, with (ΔEp)c being the displacement of the peak potential due to metal complexation, MLp, and log(M2+) is the logarithm of theoretical chemical activity. Calculations were made with the MINEQL+ program [66, 67], using stability constants reported in the literature [55, 56, 68], corresponding for each of the metals Cd, Cu, Pb, and Zn, in their aqueous free form (M2+). The parameter (ΔEp)c was calculated with eq. (1)

ΔEpC=ΔEp+RTnFlniD,MLфMLp=RTnFlnM2+bE1

where ΔEp is the observed experimental value of the displacement in half-wave potential due to complexation, i.e. the difference in half-wave potentials of the complex, MLp minus that of the free metal (in acetate), (Ep)c -(Ep)M2+, ф is the sensitivity of the determination, obtained from the linear calibration curve (not shown) of iD,M2+ vs. (M2+)std, (μA per units of chemical activity, (A uaq−1): Substitution of the common R, T, and F values, and converting ln to log, gives eq. (2):

ΔEpC=29.7×logM2+bE2

Figure 1 shows: a) at upper left, the calibration working curves to estimate the free metal activities of Zn, Cd, Pb and Cu in solution; b) at upper right, the information of the selected ligands and pH used to generate specific levels of (M2+)-values, calculated by Mineql+, for each metal–ligand system, besides the relevant parameters used for calculations of (ΔEp)C; c) at lower left, current-potential curves containing five voltammograms generated for each metal, under the predicted (M2+)-values (from left to right, respectively) for Zn, Cd, Pb, and Cu, of 9.7 × 10−6, 7.1 × 10−6, 8.2 × 10−7, 5.9 × 10−7, and 1.5 × 10−7 in 0.01 M acetic acid, pH = 4.5; and d) at lower right, voltammograms of the real samples, [M2+]ac-OM: 20% Soil:80% mine waste solution. From left to right, voltammograms correspond to Zn, Cd, Pb, and Cu, respectively. Free metal activities measured under low (5%) and high (20%) doses of compost added to mixtures (w/w) are presented in red and green, respectively. Levels of metal measured were Zn = 3 × 10−6 (red); Cd = 2.4 × 10−10 (red); Pb = 2.4 × 10−8 (red); and Cu = 3.5 × 10−15 (green) and 4.4 × 10−7 (red).

Figure 1.

a) Upper left linear graphs show (ΔEp)c-log(M2+) calibration curves for Cd, Cu, Pb, and Zn as free metals, (M2+)aq, obtained with eq. 2. (M2+) was calculated using Mineql+; b) upper right columns show calculated log(M2+)-values, measured ASV-ΔE½-values (as conditioned by ligands and pHexp) and log-(ΔEp)c graphs; c) Voltammograms at lower left, show experimental peak-current curves for Zn, Cd, Pb, and Cu, respectively, for five free Zn2+, Cd2+, Pb2+, Cu2 activities, corresponding to 9.7 × 10−6, 7.1 × 10−6, 8.2 × 10−7, 5.9 × 10−7, and 1.5 × 10−7, for each metal, in 0.01 M acetic acid pH 4.5; d) lower right voltammograms show current-potential curves for [M2+]ac-OM-treated mixture 20%-soil:80%-mine waste solutions. Peak currents in red and green represent (M2+)-values measured under low (5% w/w) and high (20%) compost doses added to mixtures. Metal activities were Zn = 3 × 10−6; Cd = 2.4 × 10−10; Pb = 2.4 × 10−8; and Cu = 4.4 × 10−7 (in red) and Cu = 3.5 × 10−15 (green). For ASV-conditions, see [55, 56].

2.2.2 The total HM soluble fraction

As mentioned before, an important step in the process of chemical speciation of pollutant metals in aqueous solutions is the analytical evaluation of the (M2+)-value [57, 58]. Although this parameter helps in the assessment and remediation of polluted sites, it is well known that a fraction of the sorbed metals may also contribute to the bioavailable fraction by replenishing into the solution, part of the exhausted ions that plants take up from the solution. Extractable fractions, i.e., the exchangeable and the readily acid-soluble precipitates (e.g., sulfates, carbonates, etc.), can also substantially contribute to the nutrition of plants, as well as a small fraction of the metals non-specifically adsorbed by organic matter and the Si, Mn, Fe and Al-oxides. Thus, although the free metal ion is the most toxic of metal species, its determination is not the only important one when evaluating the phytoavailable-toxic HM levels. Thus, the exchangeable-, acid-soluble-, and chelate-extractions, must be considered too. Thus, in the water-soluble fraction both, the (M2+)-value, and the total soluble metal must be evaluated when studying HM-polluted mine wastes.

2.3 Extraction procedures for solid phase-bound-heavy metals

One of the most widely used protocols to extract HMs sequentially was proposed by Tessier et al. [6]. Elements were separated into five “operationally” defined fractions: exchangeable; acid-soluble (carbonates); reducible (Fe/Mn oxides); oxidizable (organic matter); and residual. Other authors have referred differently to similar fractions, and even suggesting different order of sequence (chelate extractable, sulfide-associated, etc.), and even modifying concentrations, reaction times, separation procedures, etc. (BCR [7], modified BCR [8]; Geological Society of Canada (GCS)-procedure [69, 70, 71]). Based on these, diverse fractions can be visualized to include most HM-containing phases. Although many attempts to unify terms and criteria have been published, the most popular protocols and concepts will be reviewed in terms of chemical relevance to be applied to HM-polluted sites.

2.3.1 The exchangeable and chemisorbed fractions

Heavy metals extracted in the exchangeable fractions comprise both, inner and outer sphere adsorbed species. Whereas the outer sphere weakly adsorbed metal species include those retained on the solid surfaces by relatively weak electrostatic interactions (e.g., van der Waals forces) that can be released by ion-exchange processes, those metals strongly sorbed (chemisorbed and precipitated), are retained covalently by (inner-sphere-) complex interactions. Reagents used for these purposes include mostly the rather strong Mg2+ ion-exchange capacity. The most popular reagents used for these extraction procedure are MgCl2, Mg(NO3)2, CaCl2, Ca(NO3)2, KNO3, KCl, NH4Cl, CH3COONH4, CH3COOH (see Section 2.3.2), and Ba(NO3)2, among others. These reagents do not attack organic matter, silicates, or metal sulfides [6, 72], although some dissolution of carbonates has been reported [6]. Slight decrease in pH has also been reported during the extraction [73], most probably because heavy metals may displace chemisorbed-H+ ions (salt effect), or polymeric Al-ions [14] which might hydrolyze leading to a partial dissolution of carbonates and manganese oxide fractions [14, 16]. Extraction with acetate salts, particularly NH4OOCCH3, has also been used frequently in soil studies. Divalent cations, in general, are more effective than monovalent cations in ion-exchange processes, but K+ and NH4+ promote the replacement of chemisorbed metal ions in the interlayer exchange sites of some clay minerals (illite and vermiculite). Acetate ions are slightly more stable than chloro-metal-complexes reducing the readsorption and precipitation of the extracted metals and limiting pH variations because of the buffering capacity of the solution [72]. Other reagents showing similar properties have also been used, such as nitrate salts (to avoid complexation) or calcium salts (Ca2+ being sometimes more effective than Mg2+ or NH4+ in removing exchangeable ions, but showing precipitation risks with, e.g., sulfates or phosphates). Results obtained with most of these reagents have shown a good correlation with plant uptake [74]. Permanent charge sites of layer silicate clays also retain metal cations by nonspecific electrostatic forces and, in the absence of conditions that would favor metal hydrolysis (e.g., high pH), divalent (M2+) and trivalent (M3+) transition- and HM-cations show typical ion-exchange behavior on layer silicates [14]. Several studies have confirmed this for ions such as Cu2+, Co2+, Ni2+, Mn2+, etc. which retain their inner hydration sphere offering direct support for the involvement of electrostatic forces only [51]. The strength of metal bonding then, should only depend on charge, ionic radii, and hydration properties of the cation. Thus, the ionic radii series for M2+-ions seem to apply:

Ba2+>Sr2+>Ca2+>Mg2+>Hg2+>Cd2+>Zn2+>Mn2+>Fe2+>Co2+>Ni2+>Cu2+.

Based on this sequence, it results clear the usefulness of using Mg2+-ions [6] to exchange HM-divalent ions from the nonspecific adsorption sites on clay minerals and other solid phases present on soils (Fe, Mn, Al, and Si oxides and organic matter) [50]. Other ions such as K+ [75, 76, 77], NH4+ [78, 79, 80, 81], Ca2+ [78, 82], Ba2+ [76], and even H+ (from CH3COOH, [8]) have also been used with this purpose. However, chances of precipitation of Ca2+ and Ba2+ with specific anions (e.g., CO32−, PO43−, SO42−) must be considered. Increasing the concentration of K+ and NH4+ (to 1 M–2 M), lowering the concentration to 0.01 M for CaCl2, buffering pH, and adding complexing ions for Ba2+ may avoid overestimating this fraction. Transition and HMs in soils, when present at trace levels, are retained largely in non-exchangeable forms [14, 15, 16]. Schemes for complete metal extraction require extreme treatments, including the oxidative degradation of organic matter and dissolution of Fe and Mn oxides [83, 84]. Even the preferential adsorption of polymeric hydroxy-metal cations by layer silicates would not seem to account for the stability of these sorbed form of metals. Hydr-(oxides) of Si, Al, Fe, and Mn, as well as amorphous aluminosilicates offer surface sites for HM-chemisorption. According to McBride [14], evidence for the formation of surface-metal bonds includes; a) a stoichiometry of 2H+ ions released for each M2+ ion adsorbed [85]; b) a high degree of specificity shown by Al- and Fe-oxides [86], humic substances for some metals; c) changes in the surface charge properties of the oxide as a result of adsorption [38, 58], this last effect attributed to the increased surface positive charge developed by chemisorption. A generally accepted affinity series for the specific adsorption of HMs by solid phases present in soils and sediments relates directly to their increasing ability to form hydroxy complexes (metal hydrolysis). The expected order of adsorption would then be Hg > Pb > Cu> > Zn > Co > Ni > Cd [87]. Whereas this series correlates well, but not identical, with goethite and hematite, however, several authors have reported different affinity sequences [14, 50, 86, 88]. These sequences indicate that oxides and organic fractions adsorb preferentially Pb, Cu, and Zn, as compared with Cd, Ni, and Co [89, 90]. Tables 1 and 2 show the relative adsorption selectivity of solid phases for metal ions, and metal ions preferences for adsorption, respectively, if based on their chemical properties. Predicted metal affinity sequences based on their chemical properties are shown in Table 2. Table 2 shows the expected preferences of adsorption on solid phases of soils, sediments, and mine wastes, of free metal ions, based on the relevant chemical properties that could determine at first instance, the selectivity of solid phases for the metal ions. These are charge, electronegativity, ionic radii [16] which together with charge potential (z/r) or ionic potential (Z2/r), and the Pearson’s hardness parameter, Hp-value [36, 39, 91], o Softness Y-value [16, 37] (polarizability and hardness-softness), directly influence the covalent-ionic character of the adsorbed-adsorbate interaction and the relative affinity of adsorption for each metal [92].

Solid phaseAffinity sequenceReference
Amorphous Al-oxidesCu > Pb > Zn > Ni > Co > CdKinniburgh et al. [86]
Amorphous Fe-oxidesPb > Cu > Zn > Ni > Cd > CoKinniburgh et al. [86]
Goethite (FeOOH)Cu > Pb > Zn > Co > Ni > MnMcKenzie [88]
HematitePb > Cu > Zn > Co > Ni > MnMcKenzie [88]
Mn-oxide (birnesite)Pb > Cu > Mn = Co > Zn > NiMcKenzie [88]
Fulvic acid pH = 5Cu > Pb > ZnSchnitzer and Skinner [89]

Table 1.

Heavy metal affinities for some soil fractions (adapted from Ross [50]).

Chemical propertyPredicted order of affinity
ElectronegativityCu > Ni > Co > Pb > Hg > Ag > Fe > Cd > Zn > Mn > Mg > Ca > Sr.
Ionic potential (charge/radius)Ni > Mg > Cu > Co > Zn > Cd > Sr. > Pb
Ionic radiiCs+ > Rb+ > K+ > Na+ > Li+ > Ba2+ > Sr2+ > Ca2+ > Mg2+ > Hg2+ > Cd2+ > Zn2+
Pearson’s HardnessSn > Pb > Co > Ni > Fe > Mn > Zn > Cu > Cd > Hg
SoftnessPb > Cd > Co > Cu > Ni > Zn > Sr. > Mg
HydrolysisCu2+ > Pb2+ > Ni2+ > Co2+ = Zn2+ > Mg2+ > Cd2+ > Sr2+
Irving-Williams seriesBa2+ < Sr2+< Ca2+ < Mg2+ < Mn2+ < Fe2+ < Co2+ < Zn2+ < Ni2+ < Cu2+

Table 2.

Chemical properties determining metal adsorption selectivity on soils, sediments, and mine wastes solid phases [16, 37].

2.3.2 Acid-soluble fraction

The acid-soluble fraction attacks mainly acid-active solid phases, releasing HMs such as Mn and Cd, which usually co-precipitate with carbonates. This procedure attacks solid phases that become soluble at pH ≈ 5. A buffered acetic acid/acetate solution is used (0.1–1 M, pH 2–5). The HM fraction recovered under these conditions not only may come from coprecipitates with carbonate minerals but from parts of specifically adsorbed metals on clay surfaces and edges, organic matter, Fe/Mn oxyhydroxides [72], and some sulfosalts of lead, PbSO4 [93], amorphous Fe-sulfides and Fe associated with pyrrhotite [94]. This reagent releases some trace metals remaining on the specifically adsorbed sites that would other way escape the extraction in previous steps [10]. Although large proportions of total Mn are frequently found in these extracts [95], Tessier et al. [6] concluded that Fe2+ and Mn2+ were not coming from a partial attack of FeIII/MnIV oxides but from Mg/Ca carbonate coprecipitates [96, 97], and/or from Mn chemisorbed at calcite surfaces. To get a complete carbonate dissolution, a 0.5 M (pH 4.74) acetate solution can be used [98]. Complexing agents such as EDTA, are used to extract HM ions bound to organic matter too. This acid-soluble extraction procedure if used under sequential extraction protocols should be applied before the oxidation of organic matter [99].

2.3.3 The Fe and Mn hydrous oxides: The reducible fraction

Iron and Mn oxides are excellent HMs-adsorbents. By controlling the reaction Eh and pH, dissolution of metal-oxide phases can be achieved [72]. The most successful reagents to extract the total amount of metal ions associated with these oxides use both, a reducing reagent, and a ligand to retain released ions in a soluble form, the efficacy of the reagent is determined by its reduction potential and the ability to attack Fe and Mn crystalline oxyhydroxides [72]. This dissolution can take place in one to three steps, to separate amorphous and crystalline Mn and Fe-oxides. Hydroxylamine (NH2OH), oxalic acid (H2C2O4), and dithionite (Na2S2O4) are the most used reagents.

  1. Hydroxylamine (Eh° = 1.87 V) can dissolve different metal oxides, depending on pH, concentration, extracting time, and temperature. To differentiate the various Fe-oxides, warm NH2OH solutions can be used at pH 2. Acetic acid or HCl is preferred over HNO3 to avoid readsorption problems [100], taking advantage of complexing properties of ions such as Cl or CH3COO. A complete dissolution of amorphous Fe-oxides has been reported [101], skipping the attack of the crystalline phases. Other authors preferred NH2OH/CH3COOH solutions for better extraction yields than NH2OH·HCl in HNO3 [102]. Simultaneous extraction of Mn-Fe-oxides can be achieved with 0.02–0.04 M NH2OH solutions in 25% CH3COOH, at high temperatures (96–100°C). Tessier et al. [6] found total dissolution of Fe-reducible fractions within 6 h. However, the protocol seemed insufficient for Fe extraction [97, 103, 104] when Fe content is high [103, 105], for which an additional Fe-specific step is advised [97, 104, 105]. Total Mn- and amorphous Fe-oxides, and partial dissolution of crystalline Fe-oxides, can be reached at low pH (1.7) and high NH2OH·HCl concentrations (0.25 M).

  2. Oxalic acid/oxalate (H2C2O4/C2O42−) extractions are used due to the relatively high number of stable Fe-oxalate complexes (log K = 4.35–18.49 for Fe3+ and 3.20–5.15 for Fe2+) that form and the low reducing properties of the solution (E° = −0.38 V). This protocol was originally proposed to specifically eliminate amorphous Fe-oxides from other Fe solid phases in soils, following the previous Mn oxides destruction [105]. However, not only the amorphous Si- and Al-oxides are extracted (due to the very stable Al-oxalate complexes, with log K = 15) [104], but part of the crystalline Fe-oxides (maghemite [101] and lepidocrocite and goethite [106]). HMs associated with Al-oxides cannot be distinguished from those coming from the HMs-interactions with amorphous Fe (hydr-)oxides. Furthermore, as the FeII-oxalate complex auto-catalytically promotes the FeIII-reduction [107], Fe from organic complexes [101] is sometimes included too. This makes it difficult to use this extraction protocol if detailed specific speciation studies are needed. Thus, to specifically solubilize Fe (hydr-)oxides from the Fe-phases of mine wastes, a 0.2 M H2C2O4/C2O42−-solution can be used. Because the reductive dissolution of crystallized FeIII-(hydr-)oxides by oxalate is slow in the dark, the reaction can be UV-catalyzed [32], or a reducing agent, as ascorbate ions [30, 33] or dithionite [6] can be added. Ascorbic acid (HA), with variable Eh°-values (Eh° = 0.19 V for dehydro-HA, DHA/HA at pH 3.5), increases FeIII-dissolution as pH decreases, the mechanism goes through an inner-sphere complex formation of adsorbed ascorbate on the hydrous FeIII-oxide surfaces, as the electron exchange, on the FeIII/II-O bond surfaces, favors the detachment of the more labile Fe2+ from the FeIII/II-O bond surfaces [107]. Since C2O42− ions form sparingly soluble Ca and Pb salts, the extraction causes low soluble Pb-recoveries [72].

  3. Sodium dithionite (Na2S2O4, Eh° = −1.12 V), a strong reducing agent, is used to dissolve even well-crystallized Fe-oxides at pH 7–8 [6]. To avoid FeS precipitation, a strong ligand is added, and the solution is buffered to stabilize pH and redox potentials. The method [108] uses a sodium citrate-sodium dithionite solution, with additions of NaHCO3 to adjust pH to 7.3 (at 80°C). This method approximates to the combined extraction content of amorphous and crystalline Fe- oxides, but a substantial precipitation of trace metals may cause an underestimation of HMs associated with oxides due to sulfide-sulfate precipitates formation [109]. This reagent is still widely used to evaluate total reactive Fe fractions in pedology [106].

2.3.4 The organic matter and sulfides: The oxidizing fraction

Heavy metals interact in many forms, not only with organic matter, humified materials, and living organisms, in soils and sediments, but with organic detritus or sulfides of some old mine wastes deposits that may have sustained HM-hyper-accumulating plants [50, 72]. In freshly deposited mine wastes, although the content of humic substances can be limited, part of the original sulfidic material may remain, so that the levels of oxidizing fraction can be high. Under oxidizing conditions, organic materials and sulfides tend to be degraded, leading to the release of sorbed metals. So, oxidizing reagents such as H2O2 (E° = 1.77 V) or NaClO (E° = 0.90 V), and pyrophosphate ions, are frequently used in fractionation studies to extract HMs associated with organic matter and sulfidic materials. Thus, since some oxidizing agents simultaneously oxidize organic matter and sulfides, this step is more commonly named as the “oxidizing fraction”.

  1. Hydrogen peroxide, (H2O2) in dilute HNO3 is generally used to prevent metal sequestrations by the Fe hydroxide formation at high pH values. However, under these conditions, even though the oxidation is promoted by time-heating operations [72], organic matter is not completely destroyed, and sulfides are partially dissolved [6]. Maximum efficacy of OM attack was found at 3 h and 2 M H2O2 [110]). To avoid readsorption of released metals, the extraction is followed by a weak complexing with NH4OOCCH3 in HNO3 [6]. There have been controversies about the oxalate formation, as a major by-product of the organic matter destruction [72], which may cause a Fe-oxide dissolution and the precipitation of sparingly soluble oxalates. The amount of HM species extracted depends on pH, where the high levels of H+ ions exchange the non-selectively adsorbed metal cations from OM and other soil fractions. An optimal pH, however, is required to provide the best estimation of trace metals bound to soil organic matter. Because H2O2 also has reducing abilities and reduces MnO2 at pH < 5 [84], this fractionation scheme is placed after the metal-oxide extraction. Some authors prefer to apply it after Mn-oxide dissolution and before Fe-oxides dissolution because the latter step is somewhat able to extract organically complex metals. If the organic matter is oxidized after the exchangeable step to destroy the organic coating of Fe-oxide particles, facilitate the next steps.

  2. Hypochlorite, pyrophosphate, and sodium hydroxide. Because the use of sodium hypochlorite, NaClO, in alkaline conditions leads to better destruction of organic matter, besides minimizing the attack to amorphous constituents and clay minerals, it is preferred as an oxidizing reagent, instead of the H2O2 procedure. However, it must be considered that some Mn oxides are converted into MnO4 ions [72]. The 0.7 M NaClO-extraction is carried out at pH 8.5, and at high temperatures, however, since hypochlorite ions are thermo-unstable in aqueous solutions and decompose quite rapidly, it is advisable to use short extraction times. To get efficient destruction of the organic matter requires to repeat 2–3 times the oxidation step. This procedure prevents the precipitation of metal hydroxides but induces the partial destruction of carbonates. For calcareous materials, a significant part of the high carbonate content is dissolved, so that the extraction should be placed after the acid-soluble step.

  3. Sodium pyrophosphate (Na4P2O7) in basic media promotes the dispersion of organic colloids. However, at pH 10, amorphous oxides are also extracted [72]. NaOH can dissolve organic matter but also attacks aluminosilicates and clays. This last reagent leads to hydroxide precipitation and is used more for fractionation studies in sludges, which essentially have a very high content of organic matter.

2.3.5 The residual fraction

Primary and secondary minerals containing metals in the crystalline lattice constitute the bulk of this fraction. Its destruction is achieved by digestion with strong acids, such as HF, HClO4, HCl, and HNO3.

2.4 A note on sequential extraction schemes

Sequential extraction protocols are very useful experimental tools for special cases where complete characterization and HM speciation studies are required. However, in cases where a single bioavailable-toxic fraction is required, these classical extraction procedures are of less use if applied in sequence. Nevertheless, established methodologies may become more instrumental if used as single extraction methods for evaluating HM-fractions that could correlate well with plant responses when exposed to limiting or excessive concentrations of essential trace (e.g., Cu, Mn, Zn, etc.), or toxic (Pb, Cd, etc.) metals. These HM-fractionation schemes, such as those of Tessier (five steps) [6, 111], BCR (four steps) [5, 7, 112], or modified BCR (three steps) [9] serve more to evaluate the potential mobilization of metals in polluted soils, sediments, and mine wastes, where pH fluctuations, extreme potential leaching conditions, or high-risk assessment studies that might foresee floodings and other effects of severe dispersion vectors that can affect specific environments. To reduce the complexity of the procedures but maintain similar outcomes, a three-step scheme has been proposed for HM-polluted soils, sewage sludge, and for studying sulfur in soils [111, 112] and hence sulfidic mine wastes. This three-step procedure uses: acetic acid (step 1), hydroxylamine (step 2), and hydrogen peroxide (step 3). The scheme was then applied for a certification of a sediment reference material (CRM 601), and that allowed it to be validated [113]. Sequential extraction procedures are applied not without presenting several experimental and theoretical problems, mainly due to the lack of selectivity of reagents [94, 98, 114, 115], readsorption and redistribution of metals during the extraction [94, 98, 116] sample pretreatments [73, 94, 100, 103, 104, 105], and general methodological associated methods. Regarding incomplete dissolution of some phases and changes in pH can lead to controversial results regarding readsorption and the redistribution of some metals. Many authors have reported that Cu, Zn, and Pb redistribute on Fe-oxides or on humic substances [94, 117] whereas others [118] stated that redistribution was significant only for high metal contents. Carbonated species of the various metals with different solubilities KsPbCO3 = 10–13.1; KsMnCO3 = 10–9.3; and KsZnCO3 = 10−10, will show incomplete dissolution during this step, and an overestimation of the HMs extracted in the reducing fraction appear too, especially Pb [94, 105, 116] showed the lack of selectivity of these schemes toward sulfides and organic compounds. Extraction of OM by oxidative agents is also unsatisfactory because refractory compounds are not completely destroyed, and sulfides are also oxidized.

From this review, it appears that all the reagents used in the various schemes have advantages and disadvantages and there is no ideal reagent or an ideal protocol for general use. Therefore, the choice of procedure must be related to a definite objective, considering the nature of the sample.

2.5 Mining wastes and the functional extraction procedures selected

Mining of Pb-Zn-Cu ores commonly generates mine wastes rich in Pb, Zn, Cu, and Cd. Some of these tailings contain pyrite-rich materials which produce not only strong acidity when oxidized (pH values <2) but cause emissions of Zn, Pb, and Cd at levels which can cause adverse effects in terrestrial environments. It has been reported that strongly acidic Zn-rich mine wastes cause severe Zn phytotoxicity [20] and can prevent all plants from surviving on the soil. There is evidence of Zn phytotoxicity, potential Cd risk to humans if tobacco, or edible plants, are grown on contaminated soils, and Pb risk to children, if exposed to road and/or house dust [119]. Although there has been important progress in risk assessment strategies for soil metals, and research on methods to remediate Zn, Cd, and Pb polluted soils and sediments, by in situ treatments or by adding amendments (e.g., phosphates, compost, biochar, biosolids, lime-rich wood-ashes, etc.), which reverse phytotoxicity of Zn and Pb risk [11, 119], there is still a strong need to find sound methodologies to remediate HM-polluted mine wastes. The following sections will present examples of such methodologies to handle this type of HM-polluted terrestrial environment.

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3. A single extraction study of a metal-polluted mine waste of Central Mexico

3.1 Introduction

In this study, we focused mainly on lowering the bioavailability and mobility of Cd, Cu, Pb, and Zn below official environmentally safe values and to warrant a biologically clean and sustainable ecosystem. To reach this goal, two schemes were visualized. First, we assayed the addition of widely used agronomic materials consisting of lime (Ca(OH)2); gypsum (CaSO4·2H2O); P-fertilizer (KH2PO4) and compost to “treat” a gradient of soil-fresh mine tailings mixtures to assess the treatment efficacy to abate the levels of the most toxic metal species available for plant growth; in solution, the free metal ion activity, (M2+)-value, for Cd2+, Cu2+, Pb2+, and Zn2+; and on the solid phase; the so-called DTPA-phytoavailable [17], and the acid-soluble fraction imposed by some international [22, 120] standards, and a national norm [23, 24, 25, 26, 27, 28]; second, a bioassay was applied to find the conditions that allowed a sensitive indicator plant to grow in these “fertile” ameliorated media. Our studies proved to be useful in deriving soil-substrate quality criteria to establish specific strategies to verify the success of remediation processes. To evaluate the HM-toxicity abatement, both the bioavailable (acid-extractable) HM fraction and the chemical activity of the free metal ion, (M2+) were measured after incubation with the agrostabilizing treatments. Acid drainage was emulated using the standardized acetic acid extraction procedure required by norms [23, 24, 25, 26, 27, 28] and standards [22, 121].

3.2 Water, DTPA, and acid-extractable heavy metal levels

Extraction solutions to evaluate water-DTPA- and acid-extractable solutions consisted of 1) H2O-CO2 [24, 120] as saturation extract; 2) 1:2 ratio DTPA-extraction [17, 22], and 3) acetic acid (HAcO)-extraction [22], were used to obtain different species and fractions of metals from soil, mine waste, and mixtures. Extracts were analyzed for total dissolved Cd, Pb, Cu, and Zn by FAAS. Initial and in equilibrium (after incubation) extractable levels of metals, [M]HAcO, were determined at a 1:20 solid:liquid ratio in 0.2 M CH3COOH. Equilibrium-free HM ion activities, (M2+), were determined in the aqueous extracts of the treated mixtures by ASV [55, 56]. Calculations of (M2+) were carried out with MINEQL+ software [66, 67].

3.3 Experimental substrate mixtures and agrochemical treatments

Enough total mass for speciation studies and bioassays of six different substrate systems were prepared by mixing soil and mine tailings at various ratios (w/w) to emulate different degrees of soil pollution as follows: A (100:0%), B (80:20%), C (60:40%), D (40:60%), E (20:80%), and F (0:100%) soil:mine waste material. The four agrochemical treatments tested consisted in adding lime [Ca(OH)2], gypsum [CaSO4·2H2O], P-fertilizer [KH2PO4], and compost [OM] at three different doses. Agronomic materials were stoichiometrically formulated according to the initial sum of the Cd, Cu, Pb, and Zn extractabilities in 0.2 N HAcO (highest dose), DTPA (medium dose), and H2O-CO2 (lowest dose). Compost dose was added to reach 5, 10, and 20% (w/w) of OM. Blank and treated mixture systems were incubated for three weeks to reach equilibrium, adding water to keep a 1:2.5 solid:liquid ratio. For P-fertilizer, the stoichiometric addition also considered the amount of exchangeable Ca2+ levels. This test helped to discriminate treatments that efficiently decreased the HM-extractable contents from those shown by the untreated blank mixtures.

3.4 Toxicity bioassays

This bioassay was carried out only for the PO4 and OM treatments following international standard instructions (ISO 1993 [29], ISO 2005 [30]). At least seven barley (Hordeum vulgare) plants per experimental unit were grown in 100 mL black conical plastic pots (max/min radii of 3.7/1.9 cm and 5 cm height) to contain ca. 140 g of material of each of the six soil-mine waste mixtures, including the four ameliorating treatments and the three doses to give a total 72 pot systems. Root length was measured [32] and statistically analyzed with a 95–99% Fisher test of significance against the HM extracted from the mixture systems as an indicator to evaluate the efficacy of treatments and doses to lower the HM toxic effects and to assess the cleanness of the treated polluted mixtures.

3.5 Results

The presence of high amounts of Pb and Zn is common in Zimapan and they are found in combination with As, mainly in minerals of arsenopyrite (AsFeS), scorodite (FeAsO4·2H2O), and in association with pyrrhotite (Fe1-xS), pyrite and marcasite (FeS2), sphalerite (ZnS) and galena (PbS), very common minerals in the area of Zimapan [3] such that As-levels are within the reported values for this element in soils which were 19–17,384 ppm in Mexico [121, 122, 123, 124] and within 5200–40,853 ppm in mine tailings of Zimapan [3, 122, 125]. Regarding the four metals of interest levels found were within the reported values for soils 15–7200 ppm for Cu, 31–3400 ppm for Pb, and 26–6270 ppm for Zn [68, 122, 123], whereas for mine wastes in the country, our results were within the reported ranges of 186–2787 ppm, 910–9500 ppm, and 2218–32,400 for the same metals, respectively [3, 121, 122]. Official Mexican regulations [23] established levels of Cd and Pb in soils of the order of 100–300 and 3–5 mg kg−1, respectively, as hazardous to crops. These limits were not exceeded in the soil sample extracts obtained with 0.2 M HAcO. Regarding Cu and Zn, levels higher than 0.2 mg kg−1 and 1.0 mg kg−1, respectively, are reported as adequate for these micronutrients [23]. Accordingly, levels of Cd and Pb in mine tailing must not exceed 24 mg kg−1 and 120 mg kg−1, respectively, in the aqueous and/or HAcO extracts, so that Pb is not within the allowable levels when extracted with HAcO 0.2 M [26]. Zn and Cu are not potentially toxic elements regulated by Mexican official norms. The efficacy of the agronomic treatments was evaluated by comparing the initial and final quantities of the studied metals, based on the acid-extractable fraction for each experimental mixture. Figure 2 shows, in contrast with reported values, which found more than 87% decrease of the HCl-extractable concentrations of Cd, Cu, Pb, and Zn in polluted soils, after a combined CaCO3-CaHPO4 stabilizer was added [126], in our studies when agricultural lime and gypsum were applied, the [M]AcO-extracted did not show a significant HM level decrease, with respect to their initial concentrations, as compared with controls (see gypsum and lime graphs in Figure 2), moreover, redissolution process was observed for all metals except for Pb in the case of gypsum, and with some tendency to positive results for the E and F systems for Cu, Pb, and Zn. For Cd, gypsum worked well only when the soil fraction dominated (systems A, B, and C at lowest dose) but, in general, without significant differences between blank and treatments (F-test, 95%). In contrast with lime and gypsum, P-fertilizer showed excellent results (see Figure 2 at lower left) when suppressing the acid-extractable levels of Cd, Pb and Zn at any dose getting for the latter diminutions of 92% of initial quantities. For Pb the lowest dose showed a biphasic behavior indicating there exist two distinct sites for sorption which agrees with results found elsewhere [127]. For Cu, a significant decrease of [M]AcO was observed only when the dose and mine tailing contents were highest for systems C to F (F-test, 95%). Figure 2 also reveals that compost showed the best results of all amending materials where HM level suppression was more homogeneous. For Cd and Zn this treatment showed a significance reduction of the extractable metal levels at the three doses tested, although results for Zn were much more pronounced. For Cu and Pb the decrease of the extractable metal where mine tailing material was higher (systems D to F) the abatement was significant with respect to the blank system A. However, where the soil was pure or slightly polluted (systems A to C) the effect was not significant, especially for Cu-lowest and medium doses where even the metal extractability increased. For Pb, only the highest and medium doses showed some efficacy in suppressing these values. Note again the biphasic sorption for Cu and Pb at all doses, but more pronounced at the lowest one. These results completely correspond with those obtained by other authors [128, 129, 130]; who added composts, biosolids, manure, and peat materials effectively reducing Cd, Pb, and Zn mobility. These results were also consistent with the aqueous free metal, [M2+] ac, levels determined by ASV (not shown). Increment in the doses produced an important drop in the activity of this toxic chemical species even in the pure tailing systems, obtaining, in the best case, a diminution of three to five orders of magnitude orders, for example, Cu and Cd system-E treated with OM, respect to control.

Figure 2.

Effect of adding ameliorating materials lime, Ca(OH)2 (upper left), gypsum, CaSO4·2H2O (upper right), P-fertilizer, KH2PO4 (lower left) and OM-compost (lower right), at low, medium, and high doses, over the 0.2 M acid-extractable levels (mg kg−1) of Cd, Cu, Zn, and Pb (Y-axis), for six experimental mixtures soil:mine waste (s:mw): A: 100% soil, B: 80:20 s:mw, C: 60:40 s:mw, D: 40:60 s:mw, E: 20:80 s:mw, F: 100% mw. Curves show in yellow-orange the blank treatment (no ameliorating material); blue the low dose (ameliorating materials added based on the sum of the four concentrations of water-soluble metals); purple the medium dose (based on the sum of the four concentrations of the DTPA-extractable metals); and red the high dose (based on the sum of the four concentrations of the acid-extractable metals). For OM-compost, low, medium, and high doses were added to reach 5% (low dose), 10% (medium), and 20% (high dose) OM levels (w/w-basis), for the water-soluble, DTPA-extractable, and acid-extractable metals, respectively.

3.6 Biotoxicity assays

Toxicity bioassay systems (A to F) and the P-fertilizer and compost treatments were tested at the medium doses, with the only intention of evaluating if there was a chance for a positive response when applying these amending materials and the indicator-sensitive plant could prove a fertile non-toxic media was created. Root length was the agronomic parameter measured [29, 30]. Fisher test was applied to root length (95% significance) to see differences among soil-mine waste mixture treatments (OM and P-fertilizer) and doses (low, medium, and high). Figure 3 shows the results of these analyses. The effect of treatments including the null one was investigated in the six soil:mine waste systems to evaluate the effect of mine waste incorporation and predominance in these emulated scenarios of polluted soils.

Figure 3.

Histograms show the root length (cm) of barley (Hordeum vulgare) plants as affected by treatments. Upper row shows the results for the blank and compost added; lower row shows the results for blank and P-fertilizer, KH2PO4 added, both at low, medium, and high doses, respectively. Different lower-case letters mean significant differences among treatments, according to Fisher’s test of 95% significance. Low, medium, and high PO4-doses were added based on the sum of the four concentrations of the water-soluble, DTPA-, and acid-extractable metals, respectively). For OM-compost, low, medium, and high doses corresponded to 5%, 10% and 20% (w/w-basis), respectively.

Figure 3 shows that roots growing in the different mixture systems were considerably affected because of the increasing content of mine waste material added to the pristine soil, having a shortage of more than 75% of the length, when exposed to pure mine waste (blank-mixture F), respect to the pure soil (blank-mixture A). With the presence of P-fertilizer or OM treatments, a significant increment in root length with respect to control systems (PO4- and OM-mixture F) was observed, especially for the pure mine tailing mixtures (PO4-mixture F), where the highest doses improved the growth remarkably. The addition of P-fertilizer in medium and high doses was effective in providing a good media for the growth of the sensitive plant in pure mine tailing whereas compost as shown by the good response of root growth when the dose was highest. Based on these findings, it results clear the DTPA- and the acid-extractable levels of Cd, Cu, Pb and Zn, gave a good indication of the phytoavailability-phytotoxicity levels being suppressed by the ameliorating material added. PO4-medium and high doses effectively corrected most of the growing problems shown on the blank (not amended) treatment and on the low doses of PO4 added which was based on the sum of all four water-extracted metals. For the OM-compost treatments, it is important to note that low, medium, and high OM levels were chosen based on what is recommended for optimum growth of plants, according to what FAO and other OM classifications suggest which consider 5–6% as the minimum good OM-level to improve soil fertility [23].

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4. Application of sequential metal(loid) extraction to evaluate the impacted area of the Sonora River acid spill

4.1 Introduction

A spill from a copper mine dam located in northern Sonora State, Mexico, occurred on August 6, 2014. Approximately 40,000 m3 of a diluted acid solution containing iron, aluminum, copper, manganese, and minor amounts of arsenic, nickel, cadmium, zinc, lead, and chromium was released into the Sonora River basin. As part of contingency measures, the mining company built a natural gypsum dike and released 1800 tons of lime gravel along the river (190 km downstream from the source). The significant pH increase favored the precipitation of soluble ions and the sedimentation of suspended particles, mainly clays and iron oxy-hydroxides, likely with adsorbed metals [131]. These whitish sediments were removed and transported back to the mine. There was heavy rainfall the following September due to Hurricanes Norberto and Odile, which dispersed the remaining sediments along the river, possibly reaching the Rodolfo Félix Valdés Dam. However, precipitation was lower in the north, near the accident area, than in the southern portion of the basin. Additionally, pedogenic carbonates from river sediments helped raise the pH of the impacted water [132]. As a result of both natural and anthropogenic attenuation processes, in eight days the river water was already neutral in several areas [133]. The authorities consider the acid solution to have flowed 190 km downstream from the accident site to the Rodolfo Felix Valdés Dam [134]. Nevertheless, based on pH monitoring of the superficial water downstream, acidity (pH 2.6–3.7) reached no farther than 150 km downstream [135]. Total metal(loid) concentrations failed to reveal information concerning the size of the affected area, since it is naturally rich in metal(loids) and because there are waste deposits from current and historical mining in several places. Prior to the accident, significantly high metal values were found in natural compartments. The Mexican Geological Service [136] reported the geochemical composition of sediments with important metal(loid) concentrations [134]. Gonzales-Leon et al. [137] described the geological formation of the Arizpe sub-basin, reporting high natural values of several elements in the soil, including those classified as toxic to humans, such as arsenic. In 2006, the Technological Institute of Sonora reported that concentrations of arsenic, cadmium, and thallium in Sonora River basin groundwater were found to be higher than those specified in the water quality criteria for drinking water by Mexican regulations and other international institutions, such as the Environmental Protection Agency of the USA and the World Health Organization. All these data indicated high metal(loid) baselines. However, despite this fact and the important spill attenuation measures already mentioned, many residents believe that the metal(loid) concentrations found in water bodies and soils are solely due to the spill. This risk perception has caused concern and controversy among inhabitants, journalists, non-governmental organizations, and the authorities [138, 139]. Consequently, spill impact evaluation is important, but discriminating the input from metal spill from that of other sources poses a formidable challenge, since this is an area in which all the metals could have come from the same ore deposits and may share the same isotopic footprints. Total metal(loid) concentrations in sediments, their sequential extraction and bioaccessibility, and other analyses were performed to enhance information on the consequences of the acid solution spill, both for the environment and for its human inhabitants [134]. Some of these results, together with sequential extractions of selected metals in mining soil and wastes, are presented to evaluate the usefulness of chemical fractionation schemes.

4.2 Methodology

4.2.1 Sample preparation and analysis

Representative sediment samples (from impacted and non-impacted sediments in 2015), polluted soils and tailings were transported in hermetically sealed plastic containers and dried at 40°C for 48 h, ground, sieved (mesh #10, < 2 mm) and homogenized by quartering [140] Portions of 100 g were re-milled (Fritsch ball grinder), sieved (mesh #200 < 74 μm), and dried at 96°C for chemical analysis. The samples were preserved at room temperature (20°C) in hermetic containers [141]. All analyses were conducted by LABQA-UNAM (Accredited Laboratory No. R-0593-031/14 by Entidad Mexicana de Acreditación). Analytical reproducibility was inspected following the laboratory’s QA/QC analytical procedure, using spike samples and certified international reference material, and preparing blanks. All analyses were duplicates, all reagents were analytical grade or high purity, and the water was ultrapure deionized (Nanopure).

Total concentration was measured through X-ray fluorescence (XRF) with a portable model DP-6000-CC Thermo Scientific XRF Olympus analyzer, used following the 6200-method [142]. Sequential Extraction was performed using a modified Tessier et al. [6] method. The procedure consisted of five successive extractions: Fraction I (F1): Exchangeable (1 M MgCl2, pH 7, shaken 250 rpm, 1 h at room temperature (19–23°C)); Fraction II (F2): Carbonates (1 M CH3COO/CH3COOH buffer, pH 5, shaken 250 rpm, 5 h at room temperature); Fraction III (F3): Fe/Mn Oxides (0.3 M, Na2S2O4 + 0.025 M, citric acid+0.175 M, sodium citrate, shaken at 250 rpm, 5 h at 96 ± 3°C); Fraction IV (F4) Organic matter/sulfides: (3 mL [HNO3] + 5 mL 30% H2O2, pH 2, shaken 250 rpm, 5 h, at 85 ± 5°C); and Fraction V (F5): Residual phase, total concentration minus the sum of fractions I-IV. Oral gastric bio-accessibility was determined with method NOM-147 [143], analogous to SBET, RBALP, and SBRC [144]. Solid samples were mixed with glycine (C2H5NO2) at pH 1.50 ± 0.05, reached with HCl 1 M, in a ratio (1 g:100 mL); shaken 1 h at 30 rpm, controlling the temperature with an immersion recirculation heater at 37 ± 2°C. Concentrations were measured in accordance with ICP-OES EPA 6010 [145].

4.2.2 Statistical analysis

The Mann–Whitney U test was applied to test the null hypothesis in this work, that is, there will be no statistically significant differences in metal(loid) concentrations in groups of sediments, sediments and polluted soils, or sediments and tailings. This function takes two data samples as parameters, uses the median as a measure of central tendency, and then sends the results with a p-value showing the statistical significance. All analyses use a significance level of p = 0.05. If the p-value ≤0.05, the conclusion is to reject the null hypothesis and to accept a difference between the ranks of the two groups (sediments, soil, tailings).

4.3 Results and discussion

Tables 35 present the results of the sequential extraction and total concentrations of three representative metal(loids) from the acid spill: As, Cu, and Fe. No anionic sequential extraction for As was made, since the results are similar to those obtained with cationic sequential extraction [134]. The recovered fractions were: Exchangeable (I); Carbonates (II); Fe and Mn-oxide/hydroxides (III); Organic Matter/Sulfides (IV); and Residual (V). Figures 47 show the recovery percentages for the three representative elements in each fraction, and Figure 4 presents the recovery fractions in wastes and polluted soils. There are no statistically significant differences in total concentrations of As, Cu, or Fe, between impacted and baseline sediments (p values = 0.19, 0.21, and 0.07 respectively). Indeed, total As and Fe concentrations means in impacted sediments were slightly lower than those of the baseline sediments (Tables 35). On the other hand, sequential extraction does provide valuable information: As is the only element for which recovery in F3-fraction was very significant, Cu was recovered in different fractions, including F1, and Fe was retrieved from impacted sediments, predominately in fraction F5 followed by F3 (Tables 35). The dominant As fractions were F3 (linked to Mn/Fe-oxides) and F5 (residual). F3 distribution was variable along the river, with higher values in the backwaters where sediments have more easily precipitated (Figure 4). The differences between impacted and baseline sediments were only statistically significant for Cu in F1-fraction (p = 1.6 × 10−2), and for Fe in F3 (p = 3.5 × 10−3) and F5 (p = 1.4 × 10−3). Therefore, Fe can also be used as a tracer of the impact of the acid solution spill, although the differences between impacted and baseline sediments are most evident in the case of As. In samples of non-impacted sediments, fraction F5 equaled 98.95%, while in impacted sediments an important percentage of As concentration belonged to F3-fraction. Significant differences between baseline and impacted sediments were statistically proved (F3, p = 3.4 × 10−5 and F5, p = 3.2 × 10−5) (Figure 4). Significant differences were also found when statistically comparing fraction F3 of impacted sediments with the same fraction of tailings and polluted soil not affected by the spill (p = 2.5 × 10−4 and p = 1.6 × 10−3, respectively). This behavior indicates that they mainly contain arsenopyrite in relatively high concentrations, which is the most reported As-mineral in the area [146] and recovered in the F5-residual fraction [147]. After sediment 24, in samples 25 and 26, practically all of the As was recovered in F5-fraction. This behavior was also observed in the following samples (data not shown), indicating no acid solution impact downstream at those sites. The high As recovery in fraction F3 of impacted sediments is mainly a consequence of the chemical changes that took place between the waste rock and the dam (the leaching process), and likely in the river as well. The acid solution spilled from the dam was a lixiviate formed during the pretreatment of rock waste deposits. Those low-grade Cu minerals were doused with a weak sulfuric acid solution, to destroy the basic minerals occurring naturally in waste rock and that partially neutralize the added sulfuric acid. This process reduces future Cu leaching. In the waste rock, most of the As was in the form of arsenopyrite (FeAsS) and was possibly present in lower concentrations as scorodite [147]. Additionally, traces of As2O3 have been reported in the waste rock deposit. Impure sulfuric acid is added to waste rock to boost the microbiological oxidation of S-minerals. Sulfur oxidation increases Cu recuperation from Chalcocite and Chalcopyrite. Under acidic conditions, AsIII can be partially oxidized to arsenates by the MnO2 from waste rock [148]. Arsenates over pH 2 lose H+-ions, forming negatively charged species which could be retained on jarosite by sulfate substitution and/or forming inner and outer sphere complexes [149]. They can also be sorbed by schwertmannite, amorphous ferrihydrite, maghemite and goethite [150]. These As retention processes could likely happen at the dam and or in the river when basic materials added to the water lead to the formation of amorphous Fe compounds with high sorption capacity. The arsenates sorbed onto Fe compounds were recovered in F3-fraction due to FeIII-reduction. Non-oxidized AsIII (arsenopyrite) was recovered on the F5-fraction, as Ankan and Schreiber [147] also observed. From sample sites 23 and up, As was recovered in F5. Although only sediments 24 and 26 are shown in Figure 4, the same behavior was observed for the rest of the analyzed sediments (data not presented). Thus, the segment of the river impacted by the acid solution spill was no greater than 30 km, an area significantly smaller than what was initially considered before remedial action was taken (190 km), indicating that control actions were effective. Although total concentration could not be used as a guide for the impacted area, the results show that sequential extraction allowed the distinction of the impact zone from other anthropogenically polluted materials containing natural minerals. The As in fraction F3 was the best tracer.

As mg/kgSamp.S1S2S3S4S5S6S7S8S9S10S11S12S13S14S15S16S17
Tot.29.434.936.343.336.733.345.037.039.136.232.725.524.430.847.427.831.6
F1≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F2≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD0.79≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F322.625.423.029.05.35.59.96.19.89.26.74.54.15.025.018.016.2
F4≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD0.79≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F56.99.513.314.331.427.835.130.929.327.026.021.020.325.922.49.815.4
Samp.S18S19S20S21S22S23S24S25S26BL1BL2BL3BL4BL5BL6BL7BL8
Tot.3931513932242523235041353327304535
F1≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F2≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD3.09≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F325.921.433.327.422.411.6≦ LOD0.50.3≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F4≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F513.09.317.311.99.612.024.922.922.746.640.834.732.626.629.745.235.4

Table 3.

As Sequential extraction and total concentration.

LOD = Limit of detection = 0.25 mg/kg.

S = sediment, BL = base line sediment, Samp = Sample, Tot = total concentration.

Cu mg/kgSampS1S2S3S4S5S6S7S8S9S10S11S12S13S14S15S16S17
Tot146.7144.3129.7274.7279.7470.3502.0140.3136.0135.0177.7186.0136.091.7176.7170.7117.0
F11.826.621.278.0100.0229.477.6≦ LOD22.323.733.347.727.612.421.419.57.0
F212.16.34.66.511.134.216.311.87.48.211.215.010.27.92.42.14.8
F385.259.758.7151.984.2189.9226.340.744.644.654.448.144.531.451.545.666.6
F4≦ LOD≦ LOD≦ LOD1.4≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD0.1≦ LOD1.93.1
F547.651.745.236.884.516.8181.987.961.758.578.975.353.839.9101.4101.635.5
SampS18S19S20S21S22S23S24S25S26BL1BL2BL3BL4BL5BL6BL7BL8
Tot.177.3163.098.7280.7792.3298.076.7199.320.0334.0277.0269.384.0155.7125.0279.3355.0
F113.623.85.57.3≦ LOD≦ LOD6.22.30.620.65.4≦ LOD≦ LOD1.81.821.328.5
F217.58.83.98.433.925.44.74.60.622.57.4≦ LOD7.22.11.830.121.8
F3114.085.451.4254.2718.4227.821.274.310.3156.0123.043.670.060.972.1126.598.1
F49.3≦ LOD≦ LOD9.624.210.60.7≦ LOD0.57.74.8≦ LOD1.6≦ LOD1.423.017.6
F523.045.037.91.315.934.243.9118.18.1128.0136.5225.75.390.947.978.4189.0

Table 4.

Cu Sequential extraction and total concentration.

LOD = Limit of detection = 0.13 mg/kg.

S = sediment, BL = base line sediment, Samp = Sample, Tot = total concentration.

Fe mg/kgSamp.S1S2S3S4S5S6S7S8S9S10S11S12S13S14S15S16S17
Tot48,00151,651522148,977492644,90147,81151,92053974380480939,90137,5624609510841,9524082
F1≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD5≦ LOD≦ LOD≦ LOD4≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD
F25262720159543161453604002451881801568773
F310,74712,391489596732938260034473390497839574167292932574183220320,0913683
F451015≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD4≦ LOD≦ LOD4018≦ LOD
F537,24439,22428439,285182942,24744,04848,379591839736,78034,121425284920,976396
SampS18S19S20S21S22S23S24S25S26BL1BL2BL3BL4BL5BL6BL7BL8
Tot.42,35443,57563,78552,88945,41645,33641,060442847,31757,47849,63849,51749,82044,92945,30617,88810,608
F1≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD41≦ LOD5≦ LOD5≦ LOD≦ LOD135711
F2301326165≦ LOD57225≦ LOD6≦ LOD≦ LOD≦ LOD≦ LOD≦ LOD515
F37192858914,36431,97713,08472362780294282084552958292135373316303511591779
F4≦ LOD465≦ LOD≦ LOD189≦ LOD141215≦ LOD354≦ LOD9349
F535,13334,96949,38820,89132,32738,10037,993390944,47849,00546,66046,59546,24841,59542,26616,6238755

Table 5.

Fe Sequential extraction and total concentration.

LOD = Limit of detection = 0.25 mg/kg.

S = sediment, BL = base line sediment, Samp = Sample, Tot = total concentration.

Figure 4.

Sequential extraction for As.

Figure 5.

Sequential extraction for Cu.

Figure 6.

Sequential extraction for Fe.

Figure 7.

Sequential extraction for As in polluted soils and tailings.

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5. Conclusions

Based on the theoretical discussion presented, regarding the right selection of a single extraction protocol, it results clear that for functional speciation studies, it is of utmost relevance to evaluate and correlate the phytoavailable/phytotoxic HM levels determined, as affected by the ameliorating agronomic materials added, with plant growth, through the application of HM-sensitive plant bioassays.

Total concentration cannot be used as a guide to determine the area impacted by the acid spill, because no difference in total metal(loid) concentrations was noted between polluted and baseline sediments after control and remedial measures were taken. On the other hand, As fractionation is an excellent option. Two other reported strategies for the identification of areas of the river impacted by the acid solution spill are based on the detection of gypsum formed after the addition of calcite, and of jarosite formed only under the acidic conditions prevailing at the dam. However, these two methods require sophisticated equipment, as amorphous particles are practically invisible to most analytical techniques. In their stead, sequential extraction conforms to a relatively simple and inexpensive method. Monitoring metal(loid) behavior is recommended in this area to evaluate changes in sediments. The eventual attainment of an equilibrium that would form more insoluble compounds, mainly from Cu, is expected. This F1-fraction recovered metal can be desorbed, which would likely impact macro-invertebrate populations of the river.

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Acknowledgments

We thank Daniel Amaro for his general support and Reyna Roldan, Esau Jimenez, Eduardo Godoy, and Edgar Bonilla from the Laboratory of Environmental Biogeochemistry of UNAM for their valuable participation in the chemical analyses. We would also like to thank Sonia Helen Ponce Wainer for proofreading and editing.

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Written By

Arturo Aguirre Gómez and Margarita Eugenia Gutiérrez Ruiz

Submitted: 24 December 2022 Reviewed: 07 February 2023 Published: 06 March 2023