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Characterization and mobilization of toxic metals from electrolytic zinc waste
Accepted Manuscript Characterization and mobilization of toxic metals from electrolytic zinc waste M.J. Martínez Sánchez, A.M. Solano Marín, A.M. Hidalgo, C.Pérez Sirvent PII:
S0045-6535(19)31179-8
DOI:
https://doi.org/10.1016/j.chemosphere.2019.05.257
Reference:
CHEM 23992
To appear in:
ECSN
Received Date: 14 September 2018 Revised Date:
26 May 2019
Accepted Date: 28 May 2019
Please cite this article as: Sánchez, M.J.Martí., Marín, A.M.S., Hidalgo, A.M., Sirvent, C.Pé., Characterization and mobilization of toxic metals from electrolytic zinc waste, Chemosphere (2019), doi: https://doi.org/10.1016/j.chemosphere.2019.05.257. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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TITLE PAGE
CHARACTERIZATION AND MOBILIZATION OF TOXIC METALS FROM ELECTROLYTIC ZINC WASTE
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M.J. Martínez Sánchez, A.M. Solano Marín, A.M. Hidalgo, C. Pérez Sirvent.
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Department of Agricultural Chemistry, Geology and Pedology, Department of Chemical Engineering. Faculty of Chemistry. University of Murcia, E-30071, Murcia. Spain. Authors for correspondence and reprints, Corresponding author: Dr. C. Pérez-Sirvent
University of Murcia. Department of Agricultural Chemistry, Geology and Pedology.
e-mail: [email protected] Co-authors:
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Faculty of Chemistry. E-30071 Murcia. Spain.
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Dr. M.J. Martínez-Sánchez e-mail: [email protected] e-mail: [email protected]
Dr. A.M. Hidalgo
e-mail: [email protected]
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Dr. A.M. Solano-Marín.
ACCEPTED MANUSCRIPT Abstract The natural and forced mobilisation of lead, cadmium and arsenic in zinc hydrometallurgy waste is studied with the purpose of establishing potentially environmentally damaging levels and associated risks in uncontrolled situations.
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Differential X-Ray diffraction is used to study, in simulated environmental situations, the relevant role played by several mineralogical and amorphous phases. The study of potential mobility shows that all the samples considered are susceptible of releasing a significant amount of potentially toxic elements (PTEs) depending of the particular
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environmental conditions. Two situations can be considered the most problematic: the natural mobilization of the released cadmium and zinc as a result of rain, and a change
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in the redox conditions caused by an anoxic environment (flooding and/or incorporation of organic matter). The presence of massive quantities of soluble salts increases the hazard potential of these residues, mobilizing the PTEs and creating a potential
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carcinogenic risk caused by a possible oral intake for both children and adults.
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CHARACTERIZATION AND MOBILIZATION OF TOXIC METALS FROM
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ELECTROLYTIC ZINC WASTE M.J. Martínez Sánchez, A.M. Solano Marín, A.M. Hidalgo, C. Pérez Sirvent.
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Department of Agricultural Chemistry, Geology and Pedology, Department of Chemical Engineering. University of Murcia, E-30071, Murcia. Spain. e-mail: [email protected]
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1. Introduction
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Zinc is the 23rd most abundant element in the earth's crust, while sphalerite, zinc
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sulphide, remains the principal mineral ore in the world. Zinc is essential for modern
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day life and, in tonnage produced, stands fourth among all metals in world production -
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being exceeded only by iron, aluminium, and copper (Abkhoshk et al., 2014; Ma et al.,
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2011). In 2016, worldwide zinc mining production was 11.9 million tonnes according to
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the International Lead and Zinc Study Group, 2017 (www.ilzsg.org/ilzsgframe.htm). At
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present, about 75% of the world’s zinc is produced for the hydrometallurgy industry,
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mainly by acid leaching of roasted concentrated zinc sulphide, followed by electrolysis
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of the zinc sulphate solution obtained
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In hydrometallurgical zinc production (Fig. 1), wastes, which may contain the
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remains of sphalerite and other minerals that accompanied its paragenesis, are
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generated. These may also have a significant heavy and toxic metal content, besides
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zinc ferrite (ZnO·Fe2O3), precipitated Fe(OH)3 and jarosite-type compounds, goethite
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(FeO·OH) and/or hematite (Fe2O3), depending on the method used for iron removal in
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the electrolytic zinc plant, although in most methods Fe3+ precipitates as jarosite- type
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compounds (Xu et al., 2016; Raghavan et al., 1998; Peng et al., 2012). The production
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of this type of waste in the worldwide is estimated at 600,000 tonnes/year (Kangas et al,
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2017.
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According to the European Waste List, 2014/955/EU, of December 18, the waste
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of non-ferrous hydrometallurgical processes is classified by the code 11.02, and the
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sludge resulting from zinc hydrometallurgy plants (including jarosite and goethite) is
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classified as hazardous waste (11.02.02*).
ACCEPTED MANUSCRIPT It has been shown that the wastes left after zinc extraction pose potential
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environmental risks because they exhibit significant toxic and heavy metals
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solubilisation (Turan et al., 2004; Sethurajan et al., 2017). For this reason, despite the
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fact that the wastes under study were stored in closed containers or reservoirs, their
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characterisation and a study of the mobilisation of the toxic and heavy metals (As, Cd
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and Pb) they contain can be considered as providing useful information, in a simulation
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of different environmental situations, for dealing with potentially environmentally
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damaging accidents (Turan et al., 2004; Sethurajan et al., 2017; Rusen et al. 2008).
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Different research papers have studied the possibility of recovering zinc from
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industrial waste (Jha et al., 2001; Li et al. 2015), in which lead is generally the metal
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most often found (Turan et al., 2004; Sahin and Erdem, 2015), along with cadmium
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(Safarzadeh et al, 2009; Gharabaghi et al., 2012), and nickel (Gharabaghi et al., 2013).
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Precious metals, such as silver, have also been observed in Zn plant leaching wastes (Ju
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et al., 2011). In addition, a small number of authors have mentioned the presence and
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behaviour (mobility and bio-availability) of As in these wastes.
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The aim of this study was, first, to characterise the natural and forced
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mobilisation of Pb, Cd and As in zinc hydrometallurgy wastes in order to establish
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potentially environmentally damaging levels and associated risks in uncontrolled
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situations; and, secondly, to establish the importance of certain mineralogical phases in
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these wastes that simulate environmental situations using differential X-Ray diffraction.
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2. Experimental
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2.1. Materials
Thirty samples of wastes from zinc hydrometallurgical processes were obtained
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from an old production plant located near Cartagena in the SE Spain and were
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characterised by physical-chemical and mineralogical methods, ascertaining the pH,
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texture, major components and Zn, Fe, Pb, Cd and As contents. This industrial plant
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used for production the scheme given in Figure 1.
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Sediment samples were collected with a shovel in each subplot, then mixed and
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homogenized and a subsample (2 kg) was taken. After the statistical treatment of the analytical data by cluster analysis, the
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samples were assigned to seven lots, thus providing seven representative samples (T1,
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T2, T3, T4, T5, T6 and T7). Each of these was characterised and studied in mobility
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tests with different extracting agents, that providing different conditions of pH and
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redox medium. In all cases, X ray diffractograms were obtained from the residues
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remaining after each treatment and compared with those obtained from the raw sample
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using Differential X Ray Diffraction (DXRD) methodology (Pausu and Gautheyron,
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2003)
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2.2. Methods
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2.2.1 General analytical determinations
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The samples were air dried, homogenized and sieved to obtain the < 2 mm
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fraction for general analytical determinations. The pH and the electrical conductivity
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(EC) were determined in a 1:5 (weight/volume) suspension in water. A granulometric
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analysis was also carried out.
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The major components of the samples were determined by X-ray fluorescence
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spectrometry using a Philips MAGIXPro Spectrophotometer and the semiquantitative
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IQ+ program previously calibrated with certified reference materials whose matrices
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were similar to the materials being studied: NIST SRM 2711 Montana Soil, NIST SRM
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2709 San Joaquín Soil, NCS DC 73319, NCS DC 73320, NCS DC 73321, NCS DC
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73323, NCS DC 73324, NCS DC 73325, NRC PACS-1, NRC BCSS-1,
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Stock standard solutions of commercially available standard solutions (Merck)
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were used. The soluble ions present in the (1:5) water extract were quantified by ionic
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chromatography. A Metrohm ionic chromatograph (model 761 Compact IC) was used.
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The extracts were filtered through 0.45 µm nylon membrane filters to prevent any
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impurity from entering the column. To determine cations, the samples were acidified
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with 2M HNO3.
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Total content of heavy metals: The wastes were ground to a fine powder using a
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concentrated HF acid solution, 200 µl of concentrated HNO3 acid solution and 5 ml of
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water were added. When digestion was complete, the solutions were transferred to a
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volumetric flask and brought to 50 ml before measurement
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An ETHOS laboratory microwave system (Milestone, Sorisole (BG), Italy) equipped
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with temperature and pressure feedback controls, operating at a maximum exit power of
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1500W, was employed for the digestion processes. All collected wastes samples were
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submitted to four selective extractions, simulating different environmental situations.
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Water extraction: Potential Toxic Elements (PTEs) were determined in the extracts of a
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1:5 (w/v) sediment-water mixture obtained according to standards UNE-EN-12457
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(AENOR, 2003). These extracts were filtered and the supernatants were used to assess
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the natural availability of PTEs by water action and representing the soluble fraction.
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Acidic medium: For this extraction, 1g of waste sample was mixed with 50 ml of 0.1
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mol l-1 HNO3 solution. This procedure was used in order to simulate the effect of acid
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drainage on selected materials. The metals extracted in this way include those fractions
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complexed by organic matter and sorbed by or coprecipitated with hydrous oxides,
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carbonates and sulphides (Soon and Abboud, 1993).
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Complexing-reducing medium: PTE availability under anoxic conditions was evaluated
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using the dithionite-citrate extraction method (Mehra and Jackson, 1960). This
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extraction simulates the behaviour of samples in a complexing-reducing environment.
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Oxidising and acidic medium: Since wastes are exposed to air and water, an oxic
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simulation was carried out. The sample was mobilized in oxidising and acidic medium
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by adding 10 ml of concentrated hydrogen peroxide (40 % v/v) solution to 1g of sample
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(Sutherland, 2010). The pH was adjusted to 2-3, and the mixture was maintained for 1
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hour at room temperature, before heating at 85±2 °C for 1 h; a further 10 ml H2O2
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aliquot was added and then heated at 85±2 °C for 1 hour; finally, 50 ml of 1 mol l-
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1
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°C (Ruby, 1999). The extract was centrifuged and the PTE content was determined.
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Bioaccesibility: The bioaccessibility tests (Semple et al, 2004) were made according to
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ammonium acetate solution were added and the mixture was shaken for 16 h at 22±5
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soil is most likely to adhere to human hands and be ingested during hand-to-mouth
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activity. To this purpose, a gastric solution was prepared according to the standard
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operating procedure (SOP) developed by the Solubility/Bioavailability Research
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Consortium (SBRC) (Martínez-Sánchez et al., 2013; Ruby, 1999).
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Zinc, lead, cadmium and iron were determined by flame atomic absorption
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spectrometry (FAAS) using a CONTRA AA700 High Resolution Continuum Source
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Atomic Absorption Spectrophotometer. When the analyte was found in trace levels,
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lead and cadmium were determined by electrothermal atomic absorption spectrometry
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(ETAAS). Arsenic was measured by atomic fluorescence spectrometry (PSA Millenium
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Merlin 10055). The reliability of the results was verified by means of the mentioned
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certified reference materials.
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X-ray diffraction (XRD) was used to characterize the mineralogical composition
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both of the original samples and of all the residues remaining after each extraction. For
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this purpose, Cu-Kα radiation with a PW3040 Philips Diffractometer was used. The
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crystalline phases were first identified by using the Joint Committee on Powder
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Diffraction Standard, (JCPDF) database and then a semiquantitative estimation was
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carried out using the XPowder software (Martin, 2016)
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for arsenic was calculated with the equation provided by USEPA (1989):
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Exposure assessment and risk characterization. The chemical daily intake (CDI)
CDI (mg/kg day) = (Cs x IR x EF x ED x CF x FI)/ (BW x AT)
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where Cs is the concentration at the point of exposure (mg kg-1); IR is the soil ingestion
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rate (mg/day); EF the exposure frequency (day/year); ED the exposure duration (year);
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CF the unit conversion factor of 10-6 (dimensionless); FI the ingest factor
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(dimensionless); BW the body weight (kg) and AT the average time (days).
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Once the CDI values were obtained, the dimensionless carcinogenic risk (Risk),
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i.e., the incremental probability of developing cancer during a lifetime can be calculated
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by means of the simple equation (Basta, 2001): Risk = CDI x SF where SF is the so-
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unacceptable when greater than 10-5. Next, the hazard index (HI) is obtained by means
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of HI= CDI x RfD where RfD is the reference dose (units in mg/kg.day). The risk is
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considered unacceptable when HI > 1. The values of SF and RFD were obtained using
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the EPA IRIS datebase (www.epa.goviris).
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software (IBM SPSS Stastistics v-20). 3. Results
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All the statistical study of the experimental data were carried using specialized
3.1. General characteristics
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The mean values of the general analytical determinations and textural analysis
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for each group are shown in Table 1. It can be seen that, except in the cases of T1, T2
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and T3 (with acidic pH values), the samples had pH values close to neutrality. There
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was a notable presence of soluble ions, as seen from the high EC levels recorded, while
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T5 had the lowest salt content. As regards the granulometric analysis, T2, T5 and T7
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had a silty loam texture, T3 and T6 a silty clay loam texture and T1 and T4 a silty clay
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texture.
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The concentration of ions varied in a wide range, as reflected by the high
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standard deviation values. The abundance of these salts in some samples agrees with the
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high conductivity and levels of soluble ions (sulphates, chlorides and ammonium)
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found. Furthermore, the fact that these samples were taken from the surface layer
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involves an increasing number of salts caused by washing. This is very important, as the
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materials have a high concentration of soluble heavy metals.
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The presence of soluble ions in the extracts (1:5) of the wastes may be due to the
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water used in the process of zinc obtention, which had high concentrations of NO3-,
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PO43- and Mg2+ in some samples, or due to the presence of salts arising from the
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supergenic alteration of the wastes. Samples T1, T2, T3 and T4 have a high
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concentration of ammonium, which reflects its use in jarosite precipitation. Another
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cation is calcium, which comes from the gypsum present in all the samples. There is a
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close relationship between the high concentration of sulphate and chloride and soluble
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salts of the cations in the table. The main compounds and the total heavy metal contents are shown in Table 2. It
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can be seen that Zn varied in proportion from a minimum of 3% to a maximum of 13%
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and Fe from 10 to 17%. The total Pb, Cd and As contents were very high in all cases,
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the last two showing a very wide range of concentrations, as seen from the high
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standard deviations they presented.
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3.2. Mineralogical composition
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Table S.1. shows the mineralogical composition of the samples studied. The
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observed mineralogy of these wastes was not unexpected given the hydrometallurgical
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process used to obtain zinc: the sphalerite is transformed into franklinite or zinc ferrite
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ZnO·Fe2O3 during the calcination process since iron is present in the ores of zinc as an
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impurity (Claassen et al., 2002; Montanaro et al., 2001). On the other hand, the presence
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of sodium and ammonium jarosites (NaFe3(SO4)2(OH)6 and NH4Fe3(SO4)2(OH)6)
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shows that, of the three processes of iron precipitation and elimination proposed in the
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bibliography, the wastes studied are a result of a jarositic precipitation. In this process,
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iron is precipitated from an acid lixiviation solution at high temperatures (95-97 ºC) in
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the presence of Na+ or NH4+ (Ismael and Carvalho, 2003). For economic reasons, most
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jarosites are based on Na+ or NH4+, although potassium jarosite is the most stable
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(Dutrizac, 2004). The gypsum may arise from the neutralization of lixiviated acid with
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lime, carried out with the aim of eliminating sulphates. The mineralogical composition
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also showed hydrosoluble sulphates, anglesite, iron oxides and oxi-hydroxides (hematite
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and akaganeite), quartz and phyllosilicates 10 Å.
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The most characteristic sulphates of these wastes are jarosites, which are present
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in all the samples: ammonium jarosite is the most abundant phase in T1, T2 and T3, and
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natrojarosite in T5 and T6. The hydrated sulphates (HS) are represented by salts of Zn,
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Fe, Mg and NH4, with different degrees of hydration and are quite soluble in water. The
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minerals mohrite (NH4)2Fe(SO4)2·6H2O) and koktaite (NH4)2Ca(SO4)2H2O) are present
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(Mg,Mn,Zn)7(SO4)(OH)12.4H2O) and hohmannite (Fe2(SO4)2(OH)2.7H2O).The presence
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of ammonium sulphate and hexahydrated zinc shows that ammonia plays an important
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role as a precipitation agent of Fe3+ in the productive process. The minerals related to
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obtaining zinc are franklinite, gypsum and jarosites (natro and ammonium). The
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silicates and quartz observed either come from sphalerite or from sediment particles.
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Gypsum is present in all the samples at variable concentrations (6-14%).
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Furthermore, the ammonium jarosite content identifies two groups of wastes, the
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first containing high levels (T1, T2, T3 and T4), while the second group contains low
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amounts of this crystalline species and has high levels of franklinite (T5 and T6). The
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T7 sample could be included in the second group, but it contains minerals that not
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appear in the other groups such as hydrated sulphates, quartz and hematite.
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The ammonium phases identified in the first group suggest a correlation between
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these samples and the extraction processes of metallic zinc, using NH4OH as a
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precipitation agent of Fe3+ in one case and only NaOH in the other. Silicated minerals are represented by quartz and phyllosilicates of 10 Å (illite),
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whose presence in these samples is low as these minerals are not directly related to the
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industrial process. The proportions of silicates present in both groups lends weight to
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the idea that the wastes from the first group are more modern than those from the
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second, which, in old waste ponds seen to have mixed with the materials present in the
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lime-filled soils of the surroundings. This would have happened in high pH conditions,
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as seen from the fact that CaCO3 has stabilized these soils. The fact that Mg2+ is found
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in sample T7 suggests seawater was used in the industrial process.
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3.3. Differential behaviour mineralogy
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The mineralogy of the samples varied significantly with the extraction method
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(TableS.1.), the least affected and most stable species being franklinite, silicates,
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anglesite and hematite (Elikem et al, 2018). The proportion of these minerals in the
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residues of the attacked samples increased due to the total or partial dissolution of the
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rest of the crystalline and amorphous components.
ACCEPTED MANUSCRIPT The mineralogy of the wastes showed slight variations from the original sample
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after water extraction (1:5), except that the hydrated sulphates were solubilized in water,
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as was gypsum in small quantities. Hence these sulphates were not found in the residues
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of different extractions. As regards franklinite, it showed a good stability, as its levels
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increased in acid wastes and in the citrate-dithionite and oxidant medium due to the
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dissolution of gypsum, sulphates, jarosites amongst many other minerals.
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Jarosites are very stable and hard-wearing minerals because they have evolved
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in crystalline form in their formation process. This explains why jarosite was always
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present in the wastes, although the level was reduced in the face of complexing attacks.
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The jarosite content was seen to decrease after reducing-complexing medium extraction.
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The phase proportionally most attacked was ammonium jarosite, since it is less stable in
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this medium than natrojarosite. In these conditions, the factors which control the
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solubility of a given phase in an extraction medium are related to the particle size,
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degree of crystallinity and chemical properties of the compound. In this case, jarosite
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solubility in the reducing-complexing medium was caused by Fe3+ to Fe2+ reduction and
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subsequent complexation in citrate medium. Such reactions need a contact surface and
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low crystalline net with low crystallite size to facilitate reagent access to the net
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(Navarro et al., 2008). Other compounds that are attacked and dissolved in this medium
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are Fe and Al oxi-hydroxides (amorphous phases) (Pérez-Sirvent et al., 2017). These
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groups of substances may be found in variable proportions and were not identified by
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XRD.
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The soluble sulphates were generally mobilized in oxidative medium, since
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gypsum (except sample T4) and hematite residues can be appreciated in samples T1, T2
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and T3. The presence of Fe oxides could be due to oxidation reactions of Fe compounds
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or dehydration reactions of non-crystalline oxi-hydroxides.
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In acid medium (pH < 2) sulphides and carbonates were attacked, and water and
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acid soluble substances were mobilized. XRD provided no evidence of the presence of
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sulphides in the samples, but it is possible that they contained galena and pyrite.
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3.4. Simulation of mobility processes of heavy metals and arsenic.
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Figure 2, show the extraction percentages of As, Cd and Pb, respectively, in the extraction media that simulated different environmental situations. In the case of As, the percentage of water soluble element was very low or even
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null, the most effective extract being the complexing-reducing medium, which provided
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extractions of above 50% in all cases. This indicates that the As is associated with
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oxides or oxi-hydroxides of Fe and Al. An important factor is the relation between the
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pH of the samples and the efficacy of the oxidising extraction medium since the
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samples with an acid pH do not mobilise the metal in this medium since T6 and T7
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showed the highest extraction percentages. The As was not associated to soluble phases
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in water, and showed a low degree of mobilization in nitric medium in samples T1, T2,
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T3 and T5 and a moderate degree in the rest. The mobilization conditions were the same
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as for jarosite and Fe, Al and Mn oxi-hydroxides. The relationship between As and Fe
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mobilization is clear in these samples, as seen in other materials from mine wastes
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(Garcia-Lorenzo et al. 2014).
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Cadmium showed the highest degree of mobilisation in practically all the media
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used, especially in the water extraction medium, with high percentages being obtained
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in T2, T 1, T 3 and T7. The fraction of Cd extractible in acid medium was generally low
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but reached fairly high values in T2 and T1. Cd extraction in the oxidising medium was
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null except in the case of T1 and T2. The low concentration of this metal makes it
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difficult to identify any one phase. However, if the concentrations are compared in the
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different extracts, some phases can be assigned to other major metals such as Zn. In
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samples T1, T2 T3 and T5, the behavior was similar for all the extraction processes.
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The water-, acid- and oxidant medium- soluble values for Cd were very similar and
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higher than those obtained in the citrate-dithionite medium. Therefore, Cd is associated
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with soluble phases and hydrated sulphates of Zn and ammonium in proportions ranging
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from 40 to 70%. The rest is included in stable networks that are not attacked by these
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media, e.g. franklinite. These low results in the complexing-reducing medium may be
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due to a secondary precipitation reaction of soluble Cd in the form of CdS, which it
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difficult to extract.
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Samples T4, T5 and T6 did not present high levels of water-soluble Cd (5-20%),
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which agrees with the absence of soluble phases identified by XRD. The values of nitric
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acid-soluble Cd were slightly higher, and the metal reached maximum solubility in
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oxidant medium (20-30%). The behaviour of Cd in these samples was similar to that of
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Pb. This relationship lends weight to the idea that Zn, Pb and Cd monosulphides were
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present in this group of samples. Compared with Cd and As, Pb had a lower level of solubility. The oxidant
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medium provided the highest extraction values (reaching 25% in T4 and T6, for
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example), due to the fact that ammonium acetate participates in the extraction process,
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which leads to the Pb being solubilized as Pb acetate. The fraction of Pb extracted in
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acid medium was also low. The Pb soluble at acid pH is usually associated to Al, Fe and
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Mn in the form of sulphides, carbonates and oxi-hydroxides. As Pb shows a low level of
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mobilization in other media, the soluble phases which contain Pb are discarded, as is the
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Pb incorporated in jarosites. The Pb was presumably associated to anhydride sulphates
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(anglesite) and to sulphides (galene), both in proportions lower than 5% due to the fact
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that they did not appear in XRD analysis, but they did in the acid medium involving an
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oxidant agent.
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3.5. Human health risk
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Environmental exposure to arsenic and heavy metals was calculated from the
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total concentration and also from the bioaccessible content of the solid samples. The
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carcinogenic risk was calculated for children and adults using total and bioaccessible As
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values for the intake pathway. The contemplated scenario implies extreme conditions,
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with, in the worst case, a site occupation of 350 days / year, and an intake by ingestion
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of 100 mg of sediment in the case of adults and 250 mg for children.
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As can be observed from figure 3 (a, b), if the values of total As are used, in all
314
cases the calculated risk exceeds the value of 10-5, which is considered unacceptable.
315
However, using the values of bioaccessible As, the risk may be considered acceptable
316
for adults but not for children in two cases (T1 and T2).
317
Total and bioaccessible values were also used to evaluate non-carcinogenic risks
318
for As, Cd, Pb and Zn. Figure 3 (c, d) represent the non-carcinogenic risk in the samples
319
studied for adults and children.
ACCEPTED MANUSCRIPT It can be seen that the risk is higher than unity in all cases for children when
321
using total values, but can be considered acceptable for all samples in terms of
322
bioavailable As and, in some samples, bioavailable Cd and Zn. For adults, the risk is
323
acceptable only for Zn and Cd using total values, not for Pb and As. Using the
324
bioaccessible values, the non-carcinogenic risk is acceptable for all the samples.
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4. Discussion
326
Figure 4 shows the behavior of the minerals and the mobilization of PTEs that is
327
expected for each simulation. The following four scenarios are considered: (i) what is
328
happening at the present time with rainwater; (ii) what will happen in provoked
329
supergenic conditions with atmospheric oxygen and humid conditions; (iii) changes that
330
will occur in reducing conditions; (iv) the impact of leaching/leached waters.
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Rainfall dissolves water soluble minerals such as gypsum and other hydrated
332
sulphates like mohrite, boyleite, torreyite and hohmannite, present in all the samples in
333
variable quantities, releasing Cd, As and Pb, among other elements. In the obtained
334
water-extract fraction, the concentration of these elements was affected by the pH,
335
which is close to neutrality, which explains why Cd was the most mobilised element.
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In oxidising media, besides being dissolved, the sulphides and minerals
337
containing Fe+2 are transformed, increasing the acidity of the medium, which may lead
338
to the release of substantial amounts of Pb.
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The stability of the samples was most strongly affected by the reducing
340
conditions, the reduction of Fe+3 and SO42- of the jarosites favouring the formation of
341
pyrite (Campos Dos Santos et al., 2016). High levels of As associated to these minerals
342
and amorphous phases of Fe are released. Anglesite is also partly affected, its
343
proportion in the samples diminishing but never disappearing.
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344
As is well known, the alteration of materials containing sulphides of Fe
345
contributes to the acidification of the medium (Martínez-Sánchez et al., 2013), releasing
346
PTEs associated both to the soluble phases (which remain in solution) and to the
347
amorphous or more crystallised phases. These are more extreme conditions than rainfall
ACCEPTED MANUSCRIPT 348
and are conditioned by the mineralogy of the wastes/residues that are present. Calculation of the CDI for Zn, Cd, Pb and As provides high values, which
350
means an inacceptable risk (both carcinogenic and non-carcinogenic) for ingestion by
351
children in almost all cases. This situation improves considerably in the case of non-
352
carcinogenic risk if we consider only the bioaccessible values (for adults), although this
353
is not true in the case of As. In children, even if only the bioaccessible values are
354
considered, there are both carcinogenic and non-carcinogenic risks associated with the
355
oral intake. Both As and Cd strongly influence the risk in these site, even though they
356
are not among the major components, This is particularly the case with As since this is
357
the element used to calculate the carcinogenic risk (www.epa.goviris).
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The wastes resulting from Zn electrometallurgy operations share many
359
characteristics with other materials such as mining wastes but differ in their higher
360
reactivity, since they release higher amounts of PTEs and represent a greater risk. These
361
differences arise from the presence of soluble phases in these wastes that can only be
362
compared with the presence of outcrops in soils and mining wastes, where the oral risk
363
is also inacceptable (Murray et al., 2014).
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5. Conclusions
365
The study of the samples allowed the different industrial processes that occurred
366
during the hydrometallurgical production process of Zn to be identified. Two types of
367
process that influence the physicochemical characteristics of the samples can be
368
distinguished, leading to different potential behaviors and different environmental
369
impacts. The presence of massive quantities of soluble salts increases the hazard
370
potential of these residues, mobilizing the PTEs and creating a potential carcinogenic
371
risk for both children and adults.
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372
The study of potential mobility shows that all the samples should be considered
373
as a potential risk for releasing significant amounts of PTEs into the environment, As
374
mobilizable element varies according to the impact considered. Two situations can be
375
considered the most problematic: the natural mobilization of the released Cd and Zn as a
376
result of rain, and/or a change in the redox conditions caused by an anoxic environment
ACCEPTED MANUSCRIPT 377
(flooding and / or incorporation of organic matter). This situation would release a very
378
high percentage of As, and also lead to a reduction in the jarosite content. Finally, it should be noted that the compositional differences between the
380
samples related to age and the type of industrial process used are not reflected in the risk
381
acceptability calculations.
382
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6. References
Abkhoshk, E., Jorjani, E., Al-Harahsheh, M.S., Rashchi, F., Naazeri, M. 2014. Review
384
of the hydrometallurgical processing of non-sulfide zinc ores. Hydrometallurgy
385
149, 153-167.
Basta, N.T., Rodriguez, R.R., Casteel, S.W. 2001. Bioavailability and risk of arsenic
M AN U
386
SC
383
387
exposure by the soil ingestion pathway. In: Frankenberger, W.T. (Ed.),
388
Environmental Chemistry of Arsenic, Marcel Dekker Inc., New York, NY,
389
p.117.
Campos Dos Santos E., Mendonca Silva J. C., Anderson Duarte H. 2016. Pyrite
TE D
390 391
Oxidation Mechanism by Oxygen in Aqueous Medium. The Journal of Physical
392
Chemistry C , 120, 2760−2768.
Claassen, J. O., Meyer, E. H. O., Rennie, J., Sandenbergh, R. F. 2002. Iron precipitation
EP
393
from zinc-rich solutions: defining the Zincor Process. Hydrometallurgy 67, 87-
395
108.
396 397 398
AC C
394
Dutrizac, J.E. 2004. The behaviour of the rare earths during the precipitation of sodium, potassium and lead jarosites. Hydrometallurgy 73, 11-30.
Elikem, E., Laird, B.D., Hamilton, J. G., Stewart, K. J., Steven D. Siciliano, S. D. and
399
Peak D. 2018. Effects of chemical speciation on the bioaccessibility of zinc in
400
spiked and smelter-affected soils. Environmental Toxicology and Chemistry 38,
401
448-459.
402
Garcia-Lorenzo, M.L., Pérez-Sirvent, C., Molina-Ruiz, J., Martínez-Sánchez, M.J.
403
2014. Mobility indices for the assessment of metal contamination in soils
ACCEPTED MANUSCRIPT 404
affected by old mining activities. Journal of Geochemical Exploration 147, 117-
405
129.
406
Gharabaghi, M., Irannajad, M., Azadmehr, A.R. 2012. Leaching behavior of cadmium from hazardous waste. Separation and Purification Technology . 86, 9-18,
408
http://dx.doi.org/ 10.1016/j.seppur.2011.10.014.
409
RI PT
407
Gharabaghi, M., Irannajad, M., Azadmehr, A.R. 2013. Leaching kinetics of nickel
extraction from hazardous waste by sulphuric acid and optimization dissolution
411
conditions. Chemical Engineering Research & Design . Des. 91, 325-331,
412
http://dx.doi. org/10.1016/j.cherd.2012.11.016.
414
Ismael, M.R.C., Carvalho, J.M.R. 2003. Iron recovery from sulphate leach liquors in zinc hydrometallurgy. Minerals Engineering 16, 31-39.
M AN U
413
SC
410
415
Jha, M.K., Kumar, V., Singh, R.J. 2001. Review of hydrometallurgical recovery of zinc
416
from industrial wastes. Resources, Conservation and Recycling 33, 1-22.
418
419
J.C.P.D.S (Joint Committee on Powder Diffraction Standard). 1980. Mineral Powder Diffraction File. Search Manual. J.C.P.D.S. 484 pp.
TE D
417
Ju, S., Zhang, Y., Zhang, Y., Xue, P., Wang, Y. 2011. Clean hydrometallurgical route to recover zinc silver, lead, copper, cadmium and iron from hazardous jarosite
421
residues produced during zinc hydrometallurgy. Journal of Hazardous Materials
422
192, 554-558, http://dx.doi.org/10.1016/j.jhazmat.2011.05.049.
EP
420
Kangas, P., Koukkari1, P., Wilson, B.P., Lundström, M., Rastas, J., Saikkonen, P.,
424
Leppinen, J., Hintikka, V. 2017. Hydrometallrgical processing of jarosite to
425 426
AC C
423
value-added products. Paper presented at Conference in Minerals Engineering 2017, Luleå, Sweden.
427
Li, Y., Liu, H., Peng, B., Min, X., Hua, M., Peng, N., Yuang, Y., Lei, J. 2015. Study on
428
separating of zinc and iron from zinc leaching residues by roasting with
429
ammonium sulphate. Hydrometallurgy 158, 42-48.
430 431
Ma, H.W., Matsubae, K., Nakajima, K., Tsai, M.S., Shao, K.H., Chen, P.C., Lee, C.H., Nagasaka, T. 2011. Substance flow analysis of zinc cycle and current status of
ACCEPTED MANUSCRIPT 432
electric arc furnace dust management for zinc recovery in Taiwan. Resour.
433
Conservation & Recycling 56, 134-140.
434
Martin, J.D. 2016. Program for Qualitative and Quantitative Powder X-Ray Diffraction Analysis, http://www.xpowder.com/Martínez-Sánchez, M.J., Martínez-López,
436
S., Martínez-Martínez, L.B., Pérez-Sirvént, C. 2013. Importance of the oral
437
arsenic bioaccessibility factor for the characterising the risk associated with soil
438
ingestion in a mining-influenced zone. Journal of Environmental Management
439
116, 10-17.
Mehra, O. P., Jackson, M. L. 1960. Iron oxide removal from soils and clays by a
SC
440
RI PT
435
4dithionite-citrate system buffered with sodium bicarbonate. Clay Minerals
442
Bulletin. 7, 317-327.
443
M AN U
441
Montanaro, L., Bianchini, N., Rincón, J. Ma., Romero, M. 2001. Sintering behaviour of
444
pressed red mud wastes from zinc hydrometallurgy. Ceramics International 27,
445
29-37.
446
Murray, J., Kirschbaum, A., Dold, B., Mendes Guimaraes, E., Pannunzio, E. 2014. Jarosite versus soluble iron-sulfate formation and their role in acid mine
448
drainage formation at the Pan de Azúcar mine tailings (Zn-Pb-Ag), NW
449
Argentina. Minerals. 4, 477-502.
EP
TE D
447
Navarro, M.C., Pérez-Sirvent, C., Martínez-Sánchez, M.J., Vidal, J., Tovar, P.J., Bech,
451
J. 2008. Abandoned mine sites as a source of contamination by heavy metals: a
452
case study in a semi-arid zone. Journal Geochemical Exploration 96, 183-193.
453 454
455
AC C
450
Pansu, M., Gautheyrou, J. 2003. Handbook of Soil Analysis. Mineralogical, Organic and Inorganic Methods. Springer. Berlin. p.993.
Peng, N., Peng, B., Chai, L., Liu, W., Li, M., Yuan, Y., Yan, H., Hou, D.K. 2012.
456
Decomposition of zinc ferrite in zinc leaching residue by reduction roasting.
457
Procedia Environmental Sciences 16, 705-714,
458
http://dx.doi.org/10.1016/j.proenv.2012.10.097.
ACCEPTED MANUSCRIPT 459
Pérez-Sirvent, C., García-Lorenzo M.L., Hernández-Pérez C., Martínez- Sánchez. M.J.
460
2017. Assessment of potentially toxic element contamination in soils from
461
Portman Bay (SE, Spain). Journal of Soils and Sediments 1-11.
462
Raghavan, R., Mohanan, P.K., Patnaik, S.C. 1998. Innovative processing technique to produce zinc concentrate from zinc production residue with simultaneous
464
recovery of lead and silver. Hydrometallurgy 48, 225-237, http://dx.
465
doi.org/10.1016/S0304-386X (97) 00082-0.
RI PT
463
Ruby, M.V., Schoof, R., Brattin, W., Goldade, M., Post, G. Harnois, M., Mosby, D.E.,
467
Casteel, S.W., Berti, W., Carpenter, M. Edwards, D., Cragin, D. Chappell, W.,
468
1999. Advances in evaluating the oral bioavailability of inorganics in soil for use
469
of human health risk assessment. Environmental Science& Technology 33,
470
3697-3705.
M AN U
471
SC
466
Rusen¸ A., Sunkar, A.S., Topkaya, Y.A. 2008. Zinc and lead extraction from cinkur leach residues by using hydrometallurgical method. Hydrometallurgy 93 45-50,
473
http://dx.doi.org/10.1016/j.hydromet.2008.02.018.
474
TE D
472
Safarzadeh, M.S., Moradkhani, D., Ojaghi-Ilkhchi, M. 2009. Kinetics of sulfuric acid leaching of cadmium from Cd–Ni zinc plant residues. Journal of Hazardous
476
Materials 163 880-890, http://dx.doi.org/10.1016/j.jhazmat.2008.07.082.
478 479
480 481 482
Sahin, M., Erdem, M. 2015. Cleaning of high lead-bearing zinc leaching residue by recovery of lead with alkaline leaching, Hydrometallurgy 153 170-178,
AC C
477
EP
475
http://dx.doi.org/10.1016/j.hydromet.2015.03.003.
Semple KT, Doick KJ, Jones KC, Burauel P, Craven A, Harms H. 2004. Defining bioavailability and bioaccessibility of contaminated soil and sediment is complicated. Environmental Science& Technology 38, 228-231.
483
Sethurajan, M., Huguenot, D., Jain, Lens, P.N.L., R. Horn, H.A., Figueiredo, L.H.A.,
484
Van Hullebusch, E.D. 2017. Leaching and selective zinc recovery from acidic
485
leachates of zinc metallurgical leach residues. Journal of Hazardous Materials
486
324, 71-82.
ACCEPTED MANUSCRIPT 487
Soon, Y.K. and Abboud, S. 1993. Cadmium, chromium, lead and nickel. In: M.R.
488
Carter (Editor), Soil Sampling and Methods of Analysis. Canadian Society of Soil
489
Science. Lewis Publishers, Boca Raton, Florida, pp. 101-108.
491 492 493 494
Sutherland, R.A. 2010. BCR®-701: a review of 10-years of sequential extraction analyses. Anal. Chim. Acta. 680, 10-20.
RI PT
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Turan, M. D., Altundoğan, H. S., Tümen, F. 2004. Recovery of zinc and lead from zinc plant residue. Hydrometallurgy 75, 169-176.
Xu, F., Jiang, L. Li, J., Zhou, C., Wen, Y., Zhang, G., Li, Z. 2016. Mass balance and quantitative analysis of cleaner production potential in a zinc electrolysis
496
cellhouse. Journal of Cleaner Production 135, 712-720.
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Figure captions
501 502 503
Figure 1. Simplified flow sheet diagram of electrolytic process in a Spanish company (own elaboration).
504 505
Figure 2. The extraction percentages of As, Cd and Pb in the different extraction media.
507 508
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Figure 3. As carcinogenic risk from adults (a) and children (b). Non-carcinogenic risk in the samples studied from adults (c) and children (d).
509
Figure 4. Behavior of the minerals and the mobilization of PTEs for each simulation.
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Table legends
513 514 515 516
Table 1. pH (1:5), EC (1:5) (dS/m) and textural analysis (%) of the samples. Concentration (mg/l) of soluble salts in the extract (1:5). Table 2. Chemical composition.
521 522
523
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32.1
52.6
43.3
(±4.12)
(±3.51)
(±0.12) (±0.91) T2
T3
3.59
35.4
22.6
74.4
(±0.21)
(±0.82)
(±1.77)
(±3.45)
4.38
50.4
9.9
48.1
(±0.41)
(±3.08)
5
65.2
(±0.22)
(±2.33)
3.6
69.2
(±0.16)
(±3.88)
36.2
47.1
(±2.35)
(±2.51)
8.5
56.9
(±0.46)
(±2.88)
(±0.18) T4
6.50 (±0.13)
T5
6.93 (±0.08)
T6
7.02 (±0.06)
T7
Sand %
6.23 (±0.11)
(±1.03) 12.9 (±0.46) 2.88 (±0.22) 9.36 (±0.87) 27.1 (±1.03)
Na+ (mg/l)
NH4+ (mg/l)
4.2
33
7400
(±2.94)
(±3.23)
(±155.77)
33
9406
(±0.78)
(±2.32)
(±138.33)
41.9
192
14866
(±4.55)
(±4.55)
(±155.61)
29.8
25
2400
(±2.61)
(±1.79)
(±35.75)
27.3
20
13
(±1.19)
(±2.01)
3
16.7 (±1.11)
K+ (mg/l)
< q.l
< q.l
Ca2+ (mg/l)
462
(±0.41)
< q.l
Mg2+ (mg/l)
(±0.55) 47
NO3(mg/l)
PO43(mg/l)
SO42(mg/l)
767
964
(±20.55)
(±7.35)
(±1.19)
(±18.66)
(±20.4)
479
153
17
527
31
(±14.34)
(±0.06)
2366
78
319
49616
(±148.92)
(±1.11)
(±6.88)
(±301.35)
140
6
(±3.18)
(±0.08)
46
7
(±1.11)
(±0.07)
208
195
(±19.62)
(±9.42)
588
< q.l
499
(±19.99) 12
Cl(mg/l)
18
(±31.11)
< q.l
F(mg/l)
106
SC
Silt %
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3.97
EC (1:5) (dS/m)
EP
T1
pH (1:5)
686 (±20.39) 536
(±0.56) 125 (±3.66)
59 (±0.82) 21 (±0.35)
(±0.21) 2 (±0.15)
1477
(±4.46)
(±20.55)
(±0.73)
(±18.44)
(±3.22)
(±0.33)
(±10.02)
(±2.35)
3044
< q.l
733
750
(±31.12)
(±20.01)
34.6
369
144
197
637
(±2.88)
(±5.09)
(±3.98)
(±2.33)
(±26.68)
137
5
152
AC C
Sample
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Table 1. pH (1:5), EC (1:5), textural analysis of the samples and concentration of soluble salts in the extract (1:5).
(±45.55)
6
< q.l
34220 (±180.02)
< q.l
44409 (±247.11)
< q.l
12673 (±185.33)
3207 (±58.76)
< q.l
8628 (±89.89)
< q.l
50520 (±315.46)
ACCEPTED MANUSCRIPT
6.1
29.57
PbO 4.31
(±0.42) (±2.57) (±0.12) T2
4.98
29.73
4.42
Na2O 1.92 (±0.42) 1.95
MgO 0.52
Al2O3 2.1
SiO2 10.15
SO3 34.12
K2O 0.35
CaO
TiO2
4.02
l.d
(±0.12) (±0.45) (±1.35) (±3.11) (±0.02) (±0.09) 0.35
2.12
10.26
34.15
0.35
4.07
l.d
(±0.35) (±2.89) (±1.16) (±0.0.21) (±0.09) (±1.29) (±0.39) (±2.98) (±0.06) (±0.51) 4.2
(±1.32) (±2.65) (±0.20) T4
9.84
18.65
2.8
(±0.56) (±2.36) (±0.12) T5
13.57
15.41
3.02
(±1.87) (±2.67) (±0.18) T6
12.33
18.79
3.34
(±1.48) (±2.82) (±0.15) T7
15.94
12.43
2.69
(±1.75) (±2.46) (±0.10)
1.04 (±0.08) 1.58 (±0.10) 4.82 (±0.11) 3.92 (±0.15) 5.4 (±1.20)
0.62
1.45
8.92
40.02
0.21
2.28
l.d
(±0.23) (±0.71) (±0.86 (±5.32) (±0.03) (±0.13) 0.61
2.44
10.6
28.4
0.37
TE D
34.63
8.65
0.12
0.7
Cd
501
As 1523
(±0.09) (±12.7) (±54.3) 0.71
433
2085
(±0.10) (±15.1) (±65.9) 0.83
777
1874
(±0.10) (±22.3) (±34.3) 1.34
479
944
(±0.31) (±0.51) (±0.29 (±3.84) (±0.04) (±1.21) (±0.10) (±0.15) (±14.5) (±28.2) 0.59
3.65
15.06
21.06
0.71
6.94
0.24
0.35
451
1094
(±0.41) (±1.11) (±2.51 (±2.93) (±0.02) (±1.18) (±0.23) (±0.12) (±11.2) (±23.5) 0.69
EP
4.23
2.88
14.29
19.13
0.58
4.73
0.23
1.04
712
989
(±0.34) (±0.10) (±1.46) (±2.78) (±0.05) (±0.12) (±0.09) (±0.22) (±29.8) (±38.5)
AC C
T3
MnO2
SC
T1
Fe2O3
M AN U
Sample ZnO
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Table 2. Chemical composition of the samples. Major components expressed as oxides (%). Cd and As (mg.Kg-1)
3.01
1.38
7.16
25.12
0.38
1.88
0.13
2.64
967
579
(±0.51) (±0.12) (±1.27 (±3.03) (±0.04) (±0.21) (±0.05) (±0.08) (±35.7) (±4.31)
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Zn-wastes
Cd>>As>Pb
Hydrated sulphates Gypsum
OM H2O
SC
M AN U
Solubilization and oxidation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K + , H+
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O2 H2O
Reduction and complexation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, amorphous phases. Suphide precipitation (ZnS and S2Fe)
TE D
Solubilization reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, H+
Reducing conditions
EP
H2O
Supergenic conditions
Pb>>As>Cd
AC C
Rainfall
Hydrated sulphates Gypsum Sulphides minerals Natrojarosite
As>>Pb>Cd Hydrated sulphates Gypsum Amonium jarosite (Natrojarosite)
Acidic impact Impact H+ Solubilization and oxidation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, amorphous phases
Chemical mobilization
As>Pb>Cd Hydrated sulphates Gypsum Natrojarosite Sulphides minerals
Franklinite- Anglesite- Hematite- Silicated minerals
Main phases affected Non affected phases
ACCEPTED MANUSCRIPT Possible highlights
The possible carcinogenic risk posed by electrolytic zinc wastes is assessed
•
The mobility and bio-availability of arsenic in these wastes is considered
•
Special attention is paid to the mineralogical phases present in the wastes
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S0045-6535(19)31179-8
DOI:
https://doi.org/10.1016/j.chemosphere.2019.05.257
Reference:
CHEM 23992
To appear in:
ECSN
Received Date: 14 September 2018 Revised Date:
26 May 2019
Accepted Date: 28 May 2019
Please cite this article as: Sánchez, M.J.Martí., Marín, A.M.S., Hidalgo, A.M., Sirvent, C.Pé., Characterization and mobilization of toxic metals from electrolytic zinc waste, Chemosphere (2019), doi: https://doi.org/10.1016/j.chemosphere.2019.05.257. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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TITLE PAGE
CHARACTERIZATION AND MOBILIZATION OF TOXIC METALS FROM ELECTROLYTIC ZINC WASTE
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M.J. Martínez Sánchez, A.M. Solano Marín, A.M. Hidalgo, C. Pérez Sirvent.
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Department of Agricultural Chemistry, Geology and Pedology, Department of Chemical Engineering. Faculty of Chemistry. University of Murcia, E-30071, Murcia. Spain. Authors for correspondence and reprints, Corresponding author: Dr. C. Pérez-Sirvent
University of Murcia. Department of Agricultural Chemistry, Geology and Pedology.
e-mail: [email protected] Co-authors:
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Faculty of Chemistry. E-30071 Murcia. Spain.
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Dr. M.J. Martínez-Sánchez e-mail: [email protected] e-mail: [email protected]
Dr. A.M. Hidalgo
e-mail: [email protected]
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Dr. A.M. Solano-Marín.
ACCEPTED MANUSCRIPT Abstract The natural and forced mobilisation of lead, cadmium and arsenic in zinc hydrometallurgy waste is studied with the purpose of establishing potentially environmentally damaging levels and associated risks in uncontrolled situations.
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Differential X-Ray diffraction is used to study, in simulated environmental situations, the relevant role played by several mineralogical and amorphous phases. The study of potential mobility shows that all the samples considered are susceptible of releasing a significant amount of potentially toxic elements (PTEs) depending of the particular
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environmental conditions. Two situations can be considered the most problematic: the natural mobilization of the released cadmium and zinc as a result of rain, and a change
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in the redox conditions caused by an anoxic environment (flooding and/or incorporation of organic matter). The presence of massive quantities of soluble salts increases the hazard potential of these residues, mobilizing the PTEs and creating a potential
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carcinogenic risk caused by a possible oral intake for both children and adults.
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CHARACTERIZATION AND MOBILIZATION OF TOXIC METALS FROM
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ELECTROLYTIC ZINC WASTE M.J. Martínez Sánchez, A.M. Solano Marín, A.M. Hidalgo, C. Pérez Sirvent.
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Department of Agricultural Chemistry, Geology and Pedology, Department of Chemical Engineering. University of Murcia, E-30071, Murcia. Spain. e-mail: [email protected]
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1. Introduction
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Zinc is the 23rd most abundant element in the earth's crust, while sphalerite, zinc
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sulphide, remains the principal mineral ore in the world. Zinc is essential for modern
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day life and, in tonnage produced, stands fourth among all metals in world production -
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being exceeded only by iron, aluminium, and copper (Abkhoshk et al., 2014; Ma et al.,
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2011). In 2016, worldwide zinc mining production was 11.9 million tonnes according to
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the International Lead and Zinc Study Group, 2017 (www.ilzsg.org/ilzsgframe.htm). At
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present, about 75% of the world’s zinc is produced for the hydrometallurgy industry,
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mainly by acid leaching of roasted concentrated zinc sulphide, followed by electrolysis
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of the zinc sulphate solution obtained
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In hydrometallurgical zinc production (Fig. 1), wastes, which may contain the
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remains of sphalerite and other minerals that accompanied its paragenesis, are
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generated. These may also have a significant heavy and toxic metal content, besides
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zinc ferrite (ZnO·Fe2O3), precipitated Fe(OH)3 and jarosite-type compounds, goethite
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(FeO·OH) and/or hematite (Fe2O3), depending on the method used for iron removal in
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the electrolytic zinc plant, although in most methods Fe3+ precipitates as jarosite- type
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compounds (Xu et al., 2016; Raghavan et al., 1998; Peng et al., 2012). The production
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of this type of waste in the worldwide is estimated at 600,000 tonnes/year (Kangas et al,
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2017.
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According to the European Waste List, 2014/955/EU, of December 18, the waste
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of non-ferrous hydrometallurgical processes is classified by the code 11.02, and the
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sludge resulting from zinc hydrometallurgy plants (including jarosite and goethite) is
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classified as hazardous waste (11.02.02*).
ACCEPTED MANUSCRIPT It has been shown that the wastes left after zinc extraction pose potential
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environmental risks because they exhibit significant toxic and heavy metals
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solubilisation (Turan et al., 2004; Sethurajan et al., 2017). For this reason, despite the
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fact that the wastes under study were stored in closed containers or reservoirs, their
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characterisation and a study of the mobilisation of the toxic and heavy metals (As, Cd
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and Pb) they contain can be considered as providing useful information, in a simulation
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of different environmental situations, for dealing with potentially environmentally
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damaging accidents (Turan et al., 2004; Sethurajan et al., 2017; Rusen et al. 2008).
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Different research papers have studied the possibility of recovering zinc from
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industrial waste (Jha et al., 2001; Li et al. 2015), in which lead is generally the metal
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most often found (Turan et al., 2004; Sahin and Erdem, 2015), along with cadmium
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(Safarzadeh et al, 2009; Gharabaghi et al., 2012), and nickel (Gharabaghi et al., 2013).
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Precious metals, such as silver, have also been observed in Zn plant leaching wastes (Ju
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et al., 2011). In addition, a small number of authors have mentioned the presence and
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behaviour (mobility and bio-availability) of As in these wastes.
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The aim of this study was, first, to characterise the natural and forced
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mobilisation of Pb, Cd and As in zinc hydrometallurgy wastes in order to establish
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potentially environmentally damaging levels and associated risks in uncontrolled
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situations; and, secondly, to establish the importance of certain mineralogical phases in
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these wastes that simulate environmental situations using differential X-Ray diffraction.
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2. Experimental
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2.1. Materials
Thirty samples of wastes from zinc hydrometallurgical processes were obtained
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from an old production plant located near Cartagena in the SE Spain and were
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characterised by physical-chemical and mineralogical methods, ascertaining the pH,
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texture, major components and Zn, Fe, Pb, Cd and As contents. This industrial plant
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used for production the scheme given in Figure 1.
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Sediment samples were collected with a shovel in each subplot, then mixed and
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homogenized and a subsample (2 kg) was taken. After the statistical treatment of the analytical data by cluster analysis, the
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samples were assigned to seven lots, thus providing seven representative samples (T1,
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T2, T3, T4, T5, T6 and T7). Each of these was characterised and studied in mobility
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tests with different extracting agents, that providing different conditions of pH and
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redox medium. In all cases, X ray diffractograms were obtained from the residues
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remaining after each treatment and compared with those obtained from the raw sample
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using Differential X Ray Diffraction (DXRD) methodology (Pausu and Gautheyron,
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2003)
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2.2. Methods
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2.2.1 General analytical determinations
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The samples were air dried, homogenized and sieved to obtain the < 2 mm
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fraction for general analytical determinations. The pH and the electrical conductivity
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(EC) were determined in a 1:5 (weight/volume) suspension in water. A granulometric
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analysis was also carried out.
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The major components of the samples were determined by X-ray fluorescence
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spectrometry using a Philips MAGIXPro Spectrophotometer and the semiquantitative
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IQ+ program previously calibrated with certified reference materials whose matrices
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were similar to the materials being studied: NIST SRM 2711 Montana Soil, NIST SRM
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2709 San Joaquín Soil, NCS DC 73319, NCS DC 73320, NCS DC 73321, NCS DC
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73323, NCS DC 73324, NCS DC 73325, NRC PACS-1, NRC BCSS-1,
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Stock standard solutions of commercially available standard solutions (Merck)
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were used. The soluble ions present in the (1:5) water extract were quantified by ionic
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chromatography. A Metrohm ionic chromatograph (model 761 Compact IC) was used.
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The extracts were filtered through 0.45 µm nylon membrane filters to prevent any
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impurity from entering the column. To determine cations, the samples were acidified
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with 2M HNO3.
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Total content of heavy metals: The wastes were ground to a fine powder using a
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concentrated HF acid solution, 200 µl of concentrated HNO3 acid solution and 5 ml of
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water were added. When digestion was complete, the solutions were transferred to a
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volumetric flask and brought to 50 ml before measurement
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An ETHOS laboratory microwave system (Milestone, Sorisole (BG), Italy) equipped
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with temperature and pressure feedback controls, operating at a maximum exit power of
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1500W, was employed for the digestion processes. All collected wastes samples were
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submitted to four selective extractions, simulating different environmental situations.
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Water extraction: Potential Toxic Elements (PTEs) were determined in the extracts of a
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1:5 (w/v) sediment-water mixture obtained according to standards UNE-EN-12457
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(AENOR, 2003). These extracts were filtered and the supernatants were used to assess
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the natural availability of PTEs by water action and representing the soluble fraction.
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Acidic medium: For this extraction, 1g of waste sample was mixed with 50 ml of 0.1
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mol l-1 HNO3 solution. This procedure was used in order to simulate the effect of acid
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drainage on selected materials. The metals extracted in this way include those fractions
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complexed by organic matter and sorbed by or coprecipitated with hydrous oxides,
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carbonates and sulphides (Soon and Abboud, 1993).
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Complexing-reducing medium: PTE availability under anoxic conditions was evaluated
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using the dithionite-citrate extraction method (Mehra and Jackson, 1960). This
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extraction simulates the behaviour of samples in a complexing-reducing environment.
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Oxidising and acidic medium: Since wastes are exposed to air and water, an oxic
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simulation was carried out. The sample was mobilized in oxidising and acidic medium
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by adding 10 ml of concentrated hydrogen peroxide (40 % v/v) solution to 1g of sample
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(Sutherland, 2010). The pH was adjusted to 2-3, and the mixture was maintained for 1
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hour at room temperature, before heating at 85±2 °C for 1 h; a further 10 ml H2O2
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aliquot was added and then heated at 85±2 °C for 1 hour; finally, 50 ml of 1 mol l-
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1
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°C (Ruby, 1999). The extract was centrifuged and the PTE content was determined.
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Bioaccesibility: The bioaccessibility tests (Semple et al, 2004) were made according to
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ammonium acetate solution were added and the mixture was shaken for 16 h at 22±5
ACCEPTED MANUSCRIPT Martínez-Sánchez et al. (2013). The <250 mm fraction was used, since this fraction of
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soil is most likely to adhere to human hands and be ingested during hand-to-mouth
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activity. To this purpose, a gastric solution was prepared according to the standard
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operating procedure (SOP) developed by the Solubility/Bioavailability Research
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Consortium (SBRC) (Martínez-Sánchez et al., 2013; Ruby, 1999).
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Zinc, lead, cadmium and iron were determined by flame atomic absorption
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spectrometry (FAAS) using a CONTRA AA700 High Resolution Continuum Source
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Atomic Absorption Spectrophotometer. When the analyte was found in trace levels,
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lead and cadmium were determined by electrothermal atomic absorption spectrometry
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(ETAAS). Arsenic was measured by atomic fluorescence spectrometry (PSA Millenium
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Merlin 10055). The reliability of the results was verified by means of the mentioned
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certified reference materials.
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X-ray diffraction (XRD) was used to characterize the mineralogical composition
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both of the original samples and of all the residues remaining after each extraction. For
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this purpose, Cu-Kα radiation with a PW3040 Philips Diffractometer was used. The
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crystalline phases were first identified by using the Joint Committee on Powder
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Diffraction Standard, (JCPDF) database and then a semiquantitative estimation was
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carried out using the XPowder software (Martin, 2016)
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for arsenic was calculated with the equation provided by USEPA (1989):
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Exposure assessment and risk characterization. The chemical daily intake (CDI)
CDI (mg/kg day) = (Cs x IR x EF x ED x CF x FI)/ (BW x AT)
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where Cs is the concentration at the point of exposure (mg kg-1); IR is the soil ingestion
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rate (mg/day); EF the exposure frequency (day/year); ED the exposure duration (year);
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CF the unit conversion factor of 10-6 (dimensionless); FI the ingest factor
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(dimensionless); BW the body weight (kg) and AT the average time (days).
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Once the CDI values were obtained, the dimensionless carcinogenic risk (Risk),
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i.e., the incremental probability of developing cancer during a lifetime can be calculated
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by means of the simple equation (Basta, 2001): Risk = CDI x SF where SF is the so-
ACCEPTED MANUSCRIPT called cancer slope factor whose units are mg/kg.day. The risk is considered
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unacceptable when greater than 10-5. Next, the hazard index (HI) is obtained by means
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of HI= CDI x RfD where RfD is the reference dose (units in mg/kg.day). The risk is
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considered unacceptable when HI > 1. The values of SF and RFD were obtained using
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the EPA IRIS datebase (www.epa.goviris).
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software (IBM SPSS Stastistics v-20). 3. Results
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All the statistical study of the experimental data were carried using specialized
3.1. General characteristics
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The mean values of the general analytical determinations and textural analysis
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for each group are shown in Table 1. It can be seen that, except in the cases of T1, T2
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and T3 (with acidic pH values), the samples had pH values close to neutrality. There
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was a notable presence of soluble ions, as seen from the high EC levels recorded, while
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T5 had the lowest salt content. As regards the granulometric analysis, T2, T5 and T7
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had a silty loam texture, T3 and T6 a silty clay loam texture and T1 and T4 a silty clay
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texture.
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The concentration of ions varied in a wide range, as reflected by the high
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standard deviation values. The abundance of these salts in some samples agrees with the
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high conductivity and levels of soluble ions (sulphates, chlorides and ammonium)
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found. Furthermore, the fact that these samples were taken from the surface layer
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involves an increasing number of salts caused by washing. This is very important, as the
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materials have a high concentration of soluble heavy metals.
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The presence of soluble ions in the extracts (1:5) of the wastes may be due to the
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water used in the process of zinc obtention, which had high concentrations of NO3-,
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PO43- and Mg2+ in some samples, or due to the presence of salts arising from the
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supergenic alteration of the wastes. Samples T1, T2, T3 and T4 have a high
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concentration of ammonium, which reflects its use in jarosite precipitation. Another
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cation is calcium, which comes from the gypsum present in all the samples. There is a
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close relationship between the high concentration of sulphate and chloride and soluble
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salts of the cations in the table. The main compounds and the total heavy metal contents are shown in Table 2. It
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can be seen that Zn varied in proportion from a minimum of 3% to a maximum of 13%
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and Fe from 10 to 17%. The total Pb, Cd and As contents were very high in all cases,
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the last two showing a very wide range of concentrations, as seen from the high
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standard deviations they presented.
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3.2. Mineralogical composition
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Table S.1. shows the mineralogical composition of the samples studied. The
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observed mineralogy of these wastes was not unexpected given the hydrometallurgical
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process used to obtain zinc: the sphalerite is transformed into franklinite or zinc ferrite
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ZnO·Fe2O3 during the calcination process since iron is present in the ores of zinc as an
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impurity (Claassen et al., 2002; Montanaro et al., 2001). On the other hand, the presence
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of sodium and ammonium jarosites (NaFe3(SO4)2(OH)6 and NH4Fe3(SO4)2(OH)6)
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shows that, of the three processes of iron precipitation and elimination proposed in the
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bibliography, the wastes studied are a result of a jarositic precipitation. In this process,
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iron is precipitated from an acid lixiviation solution at high temperatures (95-97 ºC) in
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the presence of Na+ or NH4+ (Ismael and Carvalho, 2003). For economic reasons, most
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jarosites are based on Na+ or NH4+, although potassium jarosite is the most stable
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(Dutrizac, 2004). The gypsum may arise from the neutralization of lixiviated acid with
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lime, carried out with the aim of eliminating sulphates. The mineralogical composition
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also showed hydrosoluble sulphates, anglesite, iron oxides and oxi-hydroxides (hematite
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and akaganeite), quartz and phyllosilicates 10 Å.
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The most characteristic sulphates of these wastes are jarosites, which are present
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in all the samples: ammonium jarosite is the most abundant phase in T1, T2 and T3, and
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natrojarosite in T5 and T6. The hydrated sulphates (HS) are represented by salts of Zn,
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Fe, Mg and NH4, with different degrees of hydration and are quite soluble in water. The
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minerals mohrite (NH4)2Fe(SO4)2·6H2O) and koktaite (NH4)2Ca(SO4)2H2O) are present
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(Mg,Mn,Zn)7(SO4)(OH)12.4H2O) and hohmannite (Fe2(SO4)2(OH)2.7H2O).The presence
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of ammonium sulphate and hexahydrated zinc shows that ammonia plays an important
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role as a precipitation agent of Fe3+ in the productive process. The minerals related to
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obtaining zinc are franklinite, gypsum and jarosites (natro and ammonium). The
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silicates and quartz observed either come from sphalerite or from sediment particles.
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Gypsum is present in all the samples at variable concentrations (6-14%).
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Furthermore, the ammonium jarosite content identifies two groups of wastes, the
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first containing high levels (T1, T2, T3 and T4), while the second group contains low
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amounts of this crystalline species and has high levels of franklinite (T5 and T6). The
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T7 sample could be included in the second group, but it contains minerals that not
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appear in the other groups such as hydrated sulphates, quartz and hematite.
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The ammonium phases identified in the first group suggest a correlation between
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these samples and the extraction processes of metallic zinc, using NH4OH as a
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precipitation agent of Fe3+ in one case and only NaOH in the other. Silicated minerals are represented by quartz and phyllosilicates of 10 Å (illite),
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whose presence in these samples is low as these minerals are not directly related to the
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industrial process. The proportions of silicates present in both groups lends weight to
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the idea that the wastes from the first group are more modern than those from the
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second, which, in old waste ponds seen to have mixed with the materials present in the
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lime-filled soils of the surroundings. This would have happened in high pH conditions,
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as seen from the fact that CaCO3 has stabilized these soils. The fact that Mg2+ is found
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in sample T7 suggests seawater was used in the industrial process.
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3.3. Differential behaviour mineralogy
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The mineralogy of the samples varied significantly with the extraction method
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(TableS.1.), the least affected and most stable species being franklinite, silicates,
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anglesite and hematite (Elikem et al, 2018). The proportion of these minerals in the
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residues of the attacked samples increased due to the total or partial dissolution of the
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rest of the crystalline and amorphous components.
ACCEPTED MANUSCRIPT The mineralogy of the wastes showed slight variations from the original sample
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after water extraction (1:5), except that the hydrated sulphates were solubilized in water,
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as was gypsum in small quantities. Hence these sulphates were not found in the residues
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of different extractions. As regards franklinite, it showed a good stability, as its levels
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increased in acid wastes and in the citrate-dithionite and oxidant medium due to the
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dissolution of gypsum, sulphates, jarosites amongst many other minerals.
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Jarosites are very stable and hard-wearing minerals because they have evolved
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in crystalline form in their formation process. This explains why jarosite was always
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present in the wastes, although the level was reduced in the face of complexing attacks.
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The jarosite content was seen to decrease after reducing-complexing medium extraction.
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The phase proportionally most attacked was ammonium jarosite, since it is less stable in
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this medium than natrojarosite. In these conditions, the factors which control the
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solubility of a given phase in an extraction medium are related to the particle size,
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degree of crystallinity and chemical properties of the compound. In this case, jarosite
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solubility in the reducing-complexing medium was caused by Fe3+ to Fe2+ reduction and
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subsequent complexation in citrate medium. Such reactions need a contact surface and
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low crystalline net with low crystallite size to facilitate reagent access to the net
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(Navarro et al., 2008). Other compounds that are attacked and dissolved in this medium
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are Fe and Al oxi-hydroxides (amorphous phases) (Pérez-Sirvent et al., 2017). These
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groups of substances may be found in variable proportions and were not identified by
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XRD.
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The soluble sulphates were generally mobilized in oxidative medium, since
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gypsum (except sample T4) and hematite residues can be appreciated in samples T1, T2
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and T3. The presence of Fe oxides could be due to oxidation reactions of Fe compounds
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or dehydration reactions of non-crystalline oxi-hydroxides.
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In acid medium (pH < 2) sulphides and carbonates were attacked, and water and
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acid soluble substances were mobilized. XRD provided no evidence of the presence of
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sulphides in the samples, but it is possible that they contained galena and pyrite.
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3.4. Simulation of mobility processes of heavy metals and arsenic.
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Figure 2, show the extraction percentages of As, Cd and Pb, respectively, in the extraction media that simulated different environmental situations. In the case of As, the percentage of water soluble element was very low or even
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null, the most effective extract being the complexing-reducing medium, which provided
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extractions of above 50% in all cases. This indicates that the As is associated with
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oxides or oxi-hydroxides of Fe and Al. An important factor is the relation between the
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pH of the samples and the efficacy of the oxidising extraction medium since the
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samples with an acid pH do not mobilise the metal in this medium since T6 and T7
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showed the highest extraction percentages. The As was not associated to soluble phases
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in water, and showed a low degree of mobilization in nitric medium in samples T1, T2,
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T3 and T5 and a moderate degree in the rest. The mobilization conditions were the same
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as for jarosite and Fe, Al and Mn oxi-hydroxides. The relationship between As and Fe
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mobilization is clear in these samples, as seen in other materials from mine wastes
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(Garcia-Lorenzo et al. 2014).
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Cadmium showed the highest degree of mobilisation in practically all the media
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used, especially in the water extraction medium, with high percentages being obtained
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in T2, T 1, T 3 and T7. The fraction of Cd extractible in acid medium was generally low
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but reached fairly high values in T2 and T1. Cd extraction in the oxidising medium was
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null except in the case of T1 and T2. The low concentration of this metal makes it
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difficult to identify any one phase. However, if the concentrations are compared in the
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different extracts, some phases can be assigned to other major metals such as Zn. In
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samples T1, T2 T3 and T5, the behavior was similar for all the extraction processes.
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The water-, acid- and oxidant medium- soluble values for Cd were very similar and
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higher than those obtained in the citrate-dithionite medium. Therefore, Cd is associated
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with soluble phases and hydrated sulphates of Zn and ammonium in proportions ranging
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from 40 to 70%. The rest is included in stable networks that are not attacked by these
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media, e.g. franklinite. These low results in the complexing-reducing medium may be
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due to a secondary precipitation reaction of soluble Cd in the form of CdS, which it
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difficult to extract.
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Samples T4, T5 and T6 did not present high levels of water-soluble Cd (5-20%),
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which agrees with the absence of soluble phases identified by XRD. The values of nitric
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acid-soluble Cd were slightly higher, and the metal reached maximum solubility in
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oxidant medium (20-30%). The behaviour of Cd in these samples was similar to that of
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Pb. This relationship lends weight to the idea that Zn, Pb and Cd monosulphides were
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present in this group of samples. Compared with Cd and As, Pb had a lower level of solubility. The oxidant
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medium provided the highest extraction values (reaching 25% in T4 and T6, for
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example), due to the fact that ammonium acetate participates in the extraction process,
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which leads to the Pb being solubilized as Pb acetate. The fraction of Pb extracted in
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acid medium was also low. The Pb soluble at acid pH is usually associated to Al, Fe and
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Mn in the form of sulphides, carbonates and oxi-hydroxides. As Pb shows a low level of
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mobilization in other media, the soluble phases which contain Pb are discarded, as is the
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Pb incorporated in jarosites. The Pb was presumably associated to anhydride sulphates
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(anglesite) and to sulphides (galene), both in proportions lower than 5% due to the fact
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that they did not appear in XRD analysis, but they did in the acid medium involving an
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oxidant agent.
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3.5. Human health risk
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Environmental exposure to arsenic and heavy metals was calculated from the
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total concentration and also from the bioaccessible content of the solid samples. The
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carcinogenic risk was calculated for children and adults using total and bioaccessible As
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values for the intake pathway. The contemplated scenario implies extreme conditions,
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with, in the worst case, a site occupation of 350 days / year, and an intake by ingestion
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of 100 mg of sediment in the case of adults and 250 mg for children.
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As can be observed from figure 3 (a, b), if the values of total As are used, in all
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cases the calculated risk exceeds the value of 10-5, which is considered unacceptable.
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However, using the values of bioaccessible As, the risk may be considered acceptable
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for adults but not for children in two cases (T1 and T2).
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Total and bioaccessible values were also used to evaluate non-carcinogenic risks
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for As, Cd, Pb and Zn. Figure 3 (c, d) represent the non-carcinogenic risk in the samples
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studied for adults and children.
ACCEPTED MANUSCRIPT It can be seen that the risk is higher than unity in all cases for children when
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using total values, but can be considered acceptable for all samples in terms of
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bioavailable As and, in some samples, bioavailable Cd and Zn. For adults, the risk is
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acceptable only for Zn and Cd using total values, not for Pb and As. Using the
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bioaccessible values, the non-carcinogenic risk is acceptable for all the samples.
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4. Discussion
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Figure 4 shows the behavior of the minerals and the mobilization of PTEs that is
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expected for each simulation. The following four scenarios are considered: (i) what is
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happening at the present time with rainwater; (ii) what will happen in provoked
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supergenic conditions with atmospheric oxygen and humid conditions; (iii) changes that
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will occur in reducing conditions; (iv) the impact of leaching/leached waters.
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Rainfall dissolves water soluble minerals such as gypsum and other hydrated
332
sulphates like mohrite, boyleite, torreyite and hohmannite, present in all the samples in
333
variable quantities, releasing Cd, As and Pb, among other elements. In the obtained
334
water-extract fraction, the concentration of these elements was affected by the pH,
335
which is close to neutrality, which explains why Cd was the most mobilised element.
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In oxidising media, besides being dissolved, the sulphides and minerals
337
containing Fe+2 are transformed, increasing the acidity of the medium, which may lead
338
to the release of substantial amounts of Pb.
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336
The stability of the samples was most strongly affected by the reducing
340
conditions, the reduction of Fe+3 and SO42- of the jarosites favouring the formation of
341
pyrite (Campos Dos Santos et al., 2016). High levels of As associated to these minerals
342
and amorphous phases of Fe are released. Anglesite is also partly affected, its
343
proportion in the samples diminishing but never disappearing.
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344
As is well known, the alteration of materials containing sulphides of Fe
345
contributes to the acidification of the medium (Martínez-Sánchez et al., 2013), releasing
346
PTEs associated both to the soluble phases (which remain in solution) and to the
347
amorphous or more crystallised phases. These are more extreme conditions than rainfall
ACCEPTED MANUSCRIPT 348
and are conditioned by the mineralogy of the wastes/residues that are present. Calculation of the CDI for Zn, Cd, Pb and As provides high values, which
350
means an inacceptable risk (both carcinogenic and non-carcinogenic) for ingestion by
351
children in almost all cases. This situation improves considerably in the case of non-
352
carcinogenic risk if we consider only the bioaccessible values (for adults), although this
353
is not true in the case of As. In children, even if only the bioaccessible values are
354
considered, there are both carcinogenic and non-carcinogenic risks associated with the
355
oral intake. Both As and Cd strongly influence the risk in these site, even though they
356
are not among the major components, This is particularly the case with As since this is
357
the element used to calculate the carcinogenic risk (www.epa.goviris).
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The wastes resulting from Zn electrometallurgy operations share many
359
characteristics with other materials such as mining wastes but differ in their higher
360
reactivity, since they release higher amounts of PTEs and represent a greater risk. These
361
differences arise from the presence of soluble phases in these wastes that can only be
362
compared with the presence of outcrops in soils and mining wastes, where the oral risk
363
is also inacceptable (Murray et al., 2014).
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5. Conclusions
365
The study of the samples allowed the different industrial processes that occurred
366
during the hydrometallurgical production process of Zn to be identified. Two types of
367
process that influence the physicochemical characteristics of the samples can be
368
distinguished, leading to different potential behaviors and different environmental
369
impacts. The presence of massive quantities of soluble salts increases the hazard
370
potential of these residues, mobilizing the PTEs and creating a potential carcinogenic
371
risk for both children and adults.
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372
The study of potential mobility shows that all the samples should be considered
373
as a potential risk for releasing significant amounts of PTEs into the environment, As
374
mobilizable element varies according to the impact considered. Two situations can be
375
considered the most problematic: the natural mobilization of the released Cd and Zn as a
376
result of rain, and/or a change in the redox conditions caused by an anoxic environment
ACCEPTED MANUSCRIPT 377
(flooding and / or incorporation of organic matter). This situation would release a very
378
high percentage of As, and also lead to a reduction in the jarosite content. Finally, it should be noted that the compositional differences between the
380
samples related to age and the type of industrial process used are not reflected in the risk
381
acceptability calculations.
382
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6. References
Abkhoshk, E., Jorjani, E., Al-Harahsheh, M.S., Rashchi, F., Naazeri, M. 2014. Review
384
of the hydrometallurgical processing of non-sulfide zinc ores. Hydrometallurgy
385
149, 153-167.
Basta, N.T., Rodriguez, R.R., Casteel, S.W. 2001. Bioavailability and risk of arsenic
M AN U
386
SC
383
387
exposure by the soil ingestion pathway. In: Frankenberger, W.T. (Ed.),
388
Environmental Chemistry of Arsenic, Marcel Dekker Inc., New York, NY,
389
p.117.
Campos Dos Santos E., Mendonca Silva J. C., Anderson Duarte H. 2016. Pyrite
TE D
390 391
Oxidation Mechanism by Oxygen in Aqueous Medium. The Journal of Physical
392
Chemistry C , 120, 2760−2768.
Claassen, J. O., Meyer, E. H. O., Rennie, J., Sandenbergh, R. F. 2002. Iron precipitation
EP
393
from zinc-rich solutions: defining the Zincor Process. Hydrometallurgy 67, 87-
395
108.
396 397 398
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Dutrizac, J.E. 2004. The behaviour of the rare earths during the precipitation of sodium, potassium and lead jarosites. Hydrometallurgy 73, 11-30.
Elikem, E., Laird, B.D., Hamilton, J. G., Stewart, K. J., Steven D. Siciliano, S. D. and
399
Peak D. 2018. Effects of chemical speciation on the bioaccessibility of zinc in
400
spiked and smelter-affected soils. Environmental Toxicology and Chemistry 38,
401
448-459.
402
Garcia-Lorenzo, M.L., Pérez-Sirvent, C., Molina-Ruiz, J., Martínez-Sánchez, M.J.
403
2014. Mobility indices for the assessment of metal contamination in soils
ACCEPTED MANUSCRIPT 404
affected by old mining activities. Journal of Geochemical Exploration 147, 117-
405
129.
406
Gharabaghi, M., Irannajad, M., Azadmehr, A.R. 2012. Leaching behavior of cadmium from hazardous waste. Separation and Purification Technology . 86, 9-18,
408
http://dx.doi.org/ 10.1016/j.seppur.2011.10.014.
409
RI PT
407
Gharabaghi, M., Irannajad, M., Azadmehr, A.R. 2013. Leaching kinetics of nickel
extraction from hazardous waste by sulphuric acid and optimization dissolution
411
conditions. Chemical Engineering Research & Design . Des. 91, 325-331,
412
http://dx.doi. org/10.1016/j.cherd.2012.11.016.
414
Ismael, M.R.C., Carvalho, J.M.R. 2003. Iron recovery from sulphate leach liquors in zinc hydrometallurgy. Minerals Engineering 16, 31-39.
M AN U
413
SC
410
415
Jha, M.K., Kumar, V., Singh, R.J. 2001. Review of hydrometallurgical recovery of zinc
416
from industrial wastes. Resources, Conservation and Recycling 33, 1-22.
418
419
J.C.P.D.S (Joint Committee on Powder Diffraction Standard). 1980. Mineral Powder Diffraction File. Search Manual. J.C.P.D.S. 484 pp.
TE D
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Ju, S., Zhang, Y., Zhang, Y., Xue, P., Wang, Y. 2011. Clean hydrometallurgical route to recover zinc silver, lead, copper, cadmium and iron from hazardous jarosite
421
residues produced during zinc hydrometallurgy. Journal of Hazardous Materials
422
192, 554-558, http://dx.doi.org/10.1016/j.jhazmat.2011.05.049.
EP
420
Kangas, P., Koukkari1, P., Wilson, B.P., Lundström, M., Rastas, J., Saikkonen, P.,
424
Leppinen, J., Hintikka, V. 2017. Hydrometallrgical processing of jarosite to
425 426
AC C
423
value-added products. Paper presented at Conference in Minerals Engineering 2017, Luleå, Sweden.
427
Li, Y., Liu, H., Peng, B., Min, X., Hua, M., Peng, N., Yuang, Y., Lei, J. 2015. Study on
428
separating of zinc and iron from zinc leaching residues by roasting with
429
ammonium sulphate. Hydrometallurgy 158, 42-48.
430 431
Ma, H.W., Matsubae, K., Nakajima, K., Tsai, M.S., Shao, K.H., Chen, P.C., Lee, C.H., Nagasaka, T. 2011. Substance flow analysis of zinc cycle and current status of
ACCEPTED MANUSCRIPT 432
electric arc furnace dust management for zinc recovery in Taiwan. Resour.
433
Conservation & Recycling 56, 134-140.
434
Martin, J.D. 2016. Program for Qualitative and Quantitative Powder X-Ray Diffraction Analysis, http://www.xpowder.com/Martínez-Sánchez, M.J., Martínez-López,
436
S., Martínez-Martínez, L.B., Pérez-Sirvént, C. 2013. Importance of the oral
437
arsenic bioaccessibility factor for the characterising the risk associated with soil
438
ingestion in a mining-influenced zone. Journal of Environmental Management
439
116, 10-17.
Mehra, O. P., Jackson, M. L. 1960. Iron oxide removal from soils and clays by a
SC
440
RI PT
435
4dithionite-citrate system buffered with sodium bicarbonate. Clay Minerals
442
Bulletin. 7, 317-327.
443
M AN U
441
Montanaro, L., Bianchini, N., Rincón, J. Ma., Romero, M. 2001. Sintering behaviour of
444
pressed red mud wastes from zinc hydrometallurgy. Ceramics International 27,
445
29-37.
446
Murray, J., Kirschbaum, A., Dold, B., Mendes Guimaraes, E., Pannunzio, E. 2014. Jarosite versus soluble iron-sulfate formation and their role in acid mine
448
drainage formation at the Pan de Azúcar mine tailings (Zn-Pb-Ag), NW
449
Argentina. Minerals. 4, 477-502.
EP
TE D
447
Navarro, M.C., Pérez-Sirvent, C., Martínez-Sánchez, M.J., Vidal, J., Tovar, P.J., Bech,
451
J. 2008. Abandoned mine sites as a source of contamination by heavy metals: a
452
case study in a semi-arid zone. Journal Geochemical Exploration 96, 183-193.
453 454
455
AC C
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Pansu, M., Gautheyrou, J. 2003. Handbook of Soil Analysis. Mineralogical, Organic and Inorganic Methods. Springer. Berlin. p.993.
Peng, N., Peng, B., Chai, L., Liu, W., Li, M., Yuan, Y., Yan, H., Hou, D.K. 2012.
456
Decomposition of zinc ferrite in zinc leaching residue by reduction roasting.
457
Procedia Environmental Sciences 16, 705-714,
458
http://dx.doi.org/10.1016/j.proenv.2012.10.097.
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Pérez-Sirvent, C., García-Lorenzo M.L., Hernández-Pérez C., Martínez- Sánchez. M.J.
460
2017. Assessment of potentially toxic element contamination in soils from
461
Portman Bay (SE, Spain). Journal of Soils and Sediments 1-11.
462
Raghavan, R., Mohanan, P.K., Patnaik, S.C. 1998. Innovative processing technique to produce zinc concentrate from zinc production residue with simultaneous
464
recovery of lead and silver. Hydrometallurgy 48, 225-237, http://dx.
465
doi.org/10.1016/S0304-386X (97) 00082-0.
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Ruby, M.V., Schoof, R., Brattin, W., Goldade, M., Post, G. Harnois, M., Mosby, D.E.,
467
Casteel, S.W., Berti, W., Carpenter, M. Edwards, D., Cragin, D. Chappell, W.,
468
1999. Advances in evaluating the oral bioavailability of inorganics in soil for use
469
of human health risk assessment. Environmental Science& Technology 33,
470
3697-3705.
M AN U
471
SC
466
Rusen¸ A., Sunkar, A.S., Topkaya, Y.A. 2008. Zinc and lead extraction from cinkur leach residues by using hydrometallurgical method. Hydrometallurgy 93 45-50,
473
http://dx.doi.org/10.1016/j.hydromet.2008.02.018.
474
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472
Safarzadeh, M.S., Moradkhani, D., Ojaghi-Ilkhchi, M. 2009. Kinetics of sulfuric acid leaching of cadmium from Cd–Ni zinc plant residues. Journal of Hazardous
476
Materials 163 880-890, http://dx.doi.org/10.1016/j.jhazmat.2008.07.082.
478 479
480 481 482
Sahin, M., Erdem, M. 2015. Cleaning of high lead-bearing zinc leaching residue by recovery of lead with alkaline leaching, Hydrometallurgy 153 170-178,
AC C
477
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475
http://dx.doi.org/10.1016/j.hydromet.2015.03.003.
Semple KT, Doick KJ, Jones KC, Burauel P, Craven A, Harms H. 2004. Defining bioavailability and bioaccessibility of contaminated soil and sediment is complicated. Environmental Science& Technology 38, 228-231.
483
Sethurajan, M., Huguenot, D., Jain, Lens, P.N.L., R. Horn, H.A., Figueiredo, L.H.A.,
484
Van Hullebusch, E.D. 2017. Leaching and selective zinc recovery from acidic
485
leachates of zinc metallurgical leach residues. Journal of Hazardous Materials
486
324, 71-82.
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Soon, Y.K. and Abboud, S. 1993. Cadmium, chromium, lead and nickel. In: M.R.
488
Carter (Editor), Soil Sampling and Methods of Analysis. Canadian Society of Soil
489
Science. Lewis Publishers, Boca Raton, Florida, pp. 101-108.
491 492 493 494
Sutherland, R.A. 2010. BCR®-701: a review of 10-years of sequential extraction analyses. Anal. Chim. Acta. 680, 10-20.
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Turan, M. D., Altundoğan, H. S., Tümen, F. 2004. Recovery of zinc and lead from zinc plant residue. Hydrometallurgy 75, 169-176.
Xu, F., Jiang, L. Li, J., Zhou, C., Wen, Y., Zhang, G., Li, Z. 2016. Mass balance and quantitative analysis of cleaner production potential in a zinc electrolysis
496
cellhouse. Journal of Cleaner Production 135, 712-720.
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Figure captions
501 502 503
Figure 1. Simplified flow sheet diagram of electrolytic process in a Spanish company (own elaboration).
504 505
Figure 2. The extraction percentages of As, Cd and Pb in the different extraction media.
507 508
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Figure 3. As carcinogenic risk from adults (a) and children (b). Non-carcinogenic risk in the samples studied from adults (c) and children (d).
509
Figure 4. Behavior of the minerals and the mobilization of PTEs for each simulation.
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Table legends
513 514 515 516
Table 1. pH (1:5), EC (1:5) (dS/m) and textural analysis (%) of the samples. Concentration (mg/l) of soluble salts in the extract (1:5). Table 2. Chemical composition.
521 522
523
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32.1
52.6
43.3
(±4.12)
(±3.51)
(±0.12) (±0.91) T2
T3
3.59
35.4
22.6
74.4
(±0.21)
(±0.82)
(±1.77)
(±3.45)
4.38
50.4
9.9
48.1
(±0.41)
(±3.08)
5
65.2
(±0.22)
(±2.33)
3.6
69.2
(±0.16)
(±3.88)
36.2
47.1
(±2.35)
(±2.51)
8.5
56.9
(±0.46)
(±2.88)
(±0.18) T4
6.50 (±0.13)
T5
6.93 (±0.08)
T6
7.02 (±0.06)
T7
Sand %
6.23 (±0.11)
(±1.03) 12.9 (±0.46) 2.88 (±0.22) 9.36 (±0.87) 27.1 (±1.03)
Na+ (mg/l)
NH4+ (mg/l)
4.2
33
7400
(±2.94)
(±3.23)
(±155.77)
33
9406
(±0.78)
(±2.32)
(±138.33)
41.9
192
14866
(±4.55)
(±4.55)
(±155.61)
29.8
25
2400
(±2.61)
(±1.79)
(±35.75)
27.3
20
13
(±1.19)
(±2.01)
3
16.7 (±1.11)
K+ (mg/l)
< q.l
< q.l
Ca2+ (mg/l)
462
(±0.41)
< q.l
Mg2+ (mg/l)
(±0.55) 47
NO3(mg/l)
PO43(mg/l)
SO42(mg/l)
767
964
(±20.55)
(±7.35)
(±1.19)
(±18.66)
(±20.4)
479
153
17
527
31
(±14.34)
(±0.06)
2366
78
319
49616
(±148.92)
(±1.11)
(±6.88)
(±301.35)
140
6
(±3.18)
(±0.08)
46
7
(±1.11)
(±0.07)
208
195
(±19.62)
(±9.42)
588
< q.l
499
(±19.99) 12
Cl(mg/l)
18
(±31.11)
< q.l
F(mg/l)
106
SC
Silt %
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3.97
EC (1:5) (dS/m)
EP
T1
pH (1:5)
686 (±20.39) 536
(±0.56) 125 (±3.66)
59 (±0.82) 21 (±0.35)
(±0.21) 2 (±0.15)
1477
(±4.46)
(±20.55)
(±0.73)
(±18.44)
(±3.22)
(±0.33)
(±10.02)
(±2.35)
3044
< q.l
733
750
(±31.12)
(±20.01)
34.6
369
144
197
637
(±2.88)
(±5.09)
(±3.98)
(±2.33)
(±26.68)
137
5
152
AC C
Sample
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Table 1. pH (1:5), EC (1:5), textural analysis of the samples and concentration of soluble salts in the extract (1:5).
(±45.55)
6
< q.l
34220 (±180.02)
< q.l
44409 (±247.11)
< q.l
12673 (±185.33)
3207 (±58.76)
< q.l
8628 (±89.89)
< q.l
50520 (±315.46)
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6.1
29.57
PbO 4.31
(±0.42) (±2.57) (±0.12) T2
4.98
29.73
4.42
Na2O 1.92 (±0.42) 1.95
MgO 0.52
Al2O3 2.1
SiO2 10.15
SO3 34.12
K2O 0.35
CaO
TiO2
4.02
l.d
(±0.12) (±0.45) (±1.35) (±3.11) (±0.02) (±0.09) 0.35
2.12
10.26
34.15
0.35
4.07
l.d
(±0.35) (±2.89) (±1.16) (±0.0.21) (±0.09) (±1.29) (±0.39) (±2.98) (±0.06) (±0.51) 4.2
(±1.32) (±2.65) (±0.20) T4
9.84
18.65
2.8
(±0.56) (±2.36) (±0.12) T5
13.57
15.41
3.02
(±1.87) (±2.67) (±0.18) T6
12.33
18.79
3.34
(±1.48) (±2.82) (±0.15) T7
15.94
12.43
2.69
(±1.75) (±2.46) (±0.10)
1.04 (±0.08) 1.58 (±0.10) 4.82 (±0.11) 3.92 (±0.15) 5.4 (±1.20)
0.62
1.45
8.92
40.02
0.21
2.28
l.d
(±0.23) (±0.71) (±0.86 (±5.32) (±0.03) (±0.13) 0.61
2.44
10.6
28.4
0.37
TE D
34.63
8.65
0.12
0.7
Cd
501
As 1523
(±0.09) (±12.7) (±54.3) 0.71
433
2085
(±0.10) (±15.1) (±65.9) 0.83
777
1874
(±0.10) (±22.3) (±34.3) 1.34
479
944
(±0.31) (±0.51) (±0.29 (±3.84) (±0.04) (±1.21) (±0.10) (±0.15) (±14.5) (±28.2) 0.59
3.65
15.06
21.06
0.71
6.94
0.24
0.35
451
1094
(±0.41) (±1.11) (±2.51 (±2.93) (±0.02) (±1.18) (±0.23) (±0.12) (±11.2) (±23.5) 0.69
EP
4.23
2.88
14.29
19.13
0.58
4.73
0.23
1.04
712
989
(±0.34) (±0.10) (±1.46) (±2.78) (±0.05) (±0.12) (±0.09) (±0.22) (±29.8) (±38.5)
AC C
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MnO2
SC
T1
Fe2O3
M AN U
Sample ZnO
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Table 2. Chemical composition of the samples. Major components expressed as oxides (%). Cd and As (mg.Kg-1)
3.01
1.38
7.16
25.12
0.38
1.88
0.13
2.64
967
579
(±0.51) (±0.12) (±1.27 (±3.03) (±0.04) (±0.21) (±0.05) (±0.08) (±35.7) (±4.31)
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Zn-wastes
Cd>>As>Pb
Hydrated sulphates Gypsum
OM H2O
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Solubilization and oxidation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K + , H+
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O2 H2O
Reduction and complexation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, amorphous phases. Suphide precipitation (ZnS and S2Fe)
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Solubilization reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, H+
Reducing conditions
EP
H2O
Supergenic conditions
Pb>>As>Cd
AC C
Rainfall
Hydrated sulphates Gypsum Sulphides minerals Natrojarosite
As>>Pb>Cd Hydrated sulphates Gypsum Amonium jarosite (Natrojarosite)
Acidic impact Impact H+ Solubilization and oxidation reactions: Loss of Ca2+, SO42-, Zn2+, Cd2+, Mg2+, Na+, NH4+, Cl-, NO3-, K+, amorphous phases
Chemical mobilization
As>Pb>Cd Hydrated sulphates Gypsum Natrojarosite Sulphides minerals
Franklinite- Anglesite- Hematite- Silicated minerals
Main phases affected Non affected phases
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The possible carcinogenic risk posed by electrolytic zinc wastes is assessed
•
The mobility and bio-availability of arsenic in these wastes is considered
•
Special attention is paid to the mineralogical phases present in the wastes
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