Arsenic distribution in a pasture area impacted by past mining activities

Ecotoxicology and Environmental Safety 147 (2018) 228–237

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Arsenic distribution in a pasture area impacted by past mining activities a

a,⁎

b

c

MARK

d

P. Abad-Valle , E. Álvarez-Ayuso , A. Murciego , L.M. Muñoz-Centeno , P. Alonso-Rojo , P. Villar-Alonsoe a

Department of Environmental Geochemistry, IRNASA (CSIC), C/ Cordel de Merinas 40-52, 37008 Salamanca, Spain Department of Geology, Salamanca University, Plza. de los Caídos s/n, 37008 Salamanca, Spain Department of Botany, Salamanca University, Avda. Ldo. Méndez Nieto s/n, 37007 Salamanca, Spain d Department of Edaphology, Salamanca University, Avda. Filiberto Villalobos 117, 37007 Salamanca, Spain e Saloro SLU, Avda. Italia 8, 37006 Salamanca, Spain b c

A R T I C L E I N F O

A B S T R A C T

Keywords: Arsenic Grazing land Soil pollution Phytoavailability Risk assessment

Former mine exploitations entail a serious threat to surrounding ecosystems as after closure of mining activities their unmanaged wastes can be a continuous source of toxic trace elements. Quite often these mine sites are found within agricultural farming areas, involving serious hazards as regards product (feed/food) quality. In this work a grazing land impacted by the abandoned mine exploitation of an arsenical deposit was studied so as to evaluate the fate of arsenic (As) and other trace elements and the potential risks involved. With this aim, profile soil samples (0–50 cm) and pasture plant species (Agrostis truncatula, Holcus annus and Leontodon longirostris) were collected at different distances (0–100 m) from the mine waste dump and analyzed for their trace element content and distribution. Likewise, plant trace element accumulation from impacted grazing soils and plant trace element translocation were assessed. The exposure of livestock grazing animals to As was also evaluated, establishing its acceptability regarding food safety and animal health. International soil guideline values for As in grazing land soils (50 mg kg−1) resulted greatly exceeded (up to about 20-fold) in the studied mining-affected soils. Moreover, As showed a high mobilization potential under circumstances such as phosphate application or establishment of reducing conditions. Arsenic exhibited relatively high translocation factor (TF) values (up to 0.32–0.89) in pasture plant species, reaching unsafe concentrations in their above-ground tissues (up to 32.9, 16.9 and 9.0 mg kg−1 in Agrostis truncatula, Leontodon longirostris and Holcus annus, respectively). Such concentrations represent an elevated risk of As transfer to the high trophic-chain levels as established by international legislation. The limited fraction of arsenite found in plant roots should play an important role in the relatively high As root-to-shoot translocation shown by these plant species. Both soil ingestion and pasture intake resulted important entrance pathways of As into livestock animals, showing quite close contribution levels. The cow acceptable daily intake (ADI) of As regarding food safety was surpassed in some locations of the study area when the species Agrostis truncatula was considered as the only pasture feed. Restrictions in the grazing use of lands with considerable As contents where this plant was the predominant pasture species should be established in order to preserve food quality. Therefore, the exposure of livestock animals to As via both soil ingestion and pasture consumption should be taken into account to establish the suitability of mining-impacted areas for gazing.

1. Introduction Arsenic (As) occurs naturally in all the environmental compartments (Mandal and Suzuki, 2002). Typical As concentrations in noncontaminated soils range between 1 and 40 mg kg−1, depending on the parent material (Kabata-Pendias and Pendias, 1992), but rarely exceed 10 mg kg−1 (Adriano, 1986). However, such restricted levels can be importantly increased in areas impacted by certain human activities.

⁎

Primary anthropogenic As sources are mining activities, agricultural practices, such as the use of As-contaminated irrigation water, Ascontaining pesticides and herbicides, phosphate fertilizers, sewage sludge and manure, and industrial processes, including manufacture of glass/ceramics, alloys, electronics and pigments, combustion of fossil fuels for energy production and smelting (Alloway, 1995; KabataPendias and Mukherjee, 2007). Arsenic is present as a major constituent in over 200 minerals

Corresponding author. E-mail address: [email protected] (E. Álvarez-Ayuso).

http://dx.doi.org/10.1016/j.ecoenv.2017.08.031 Received 19 December 2016; Received in revised form 11 August 2017; Accepted 14 August 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.

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uncommon (Rodrigues et al., 2012; Martínez-López et al., 2014). Furthermore, the soil physicochemical characteristics also play an important role on the soil-plant transfer of As, with pH, redox potential, phosphate content and iron and aluminum oxide content being highly influential (Adriano, 1986; Moreno-Jiménez et al., 2012). Thus, under oxic conditions the mobility of As is fairly low in acid soils with high metal oxide content (McBride, 1994), although it could be greatly increased by the presence of phosphate as this anion competes very effectively with arsenate for soil adsorption sites (Manning and Goldberg, 1996). Nevertheless, both also compete for their uptake by plant roots via phosphate transporters for which phosphate shows a higher affinity than arsenate (Tripathi et al., 2007; Zhao et al., 2009). Therefore, numerous factors can affect the soil-plant-animal transfer of As. The environmental characterization of As-polluted areas and the identification of the main factors responsible for the exposition of animals and/or humans to this toxic element are crucial to establish the suitable management options for this kind of scenarios in order to preserve product (feed/food) quality and protect public health. The main goals of this work were to 1) perform the soil environmental characterization of a pasture area affected by the former mine exploitation of an arsenical tungsten deposit, 2) assess the accumulation of As and other trace elements by pasture plant species of this area, 3) determine their suitability for feeding grazing animals according to international legislation, and 4) evaluate the exposure of grazing animals to As via both soil ingestion and pasture consumption, establishing its acceptability regarding food safety and animal health.

(Adriano, 1986), with arsenopyrite (FeAsS) being the most common one (Alloway, 1995). It is found in a wide variety of mineral deposits, in close association with many elements such as copper (Cu), lead (Pb), zinc (Zn), silver (Ag), gold (Au), tungsten (W), tin (Sn), nickel (Ni) and cobalt (Co) (Alloway, 1995; Smedley and Kinniburgh, 2002). Mining and beneficiation of these resources have generated large amounts of As-rich mine wastes, especially at historic mine sites. The transport by either wind or water of these wastes and the unconstrained leaching from them constitute a significant source of As pollution in the surrounding area. Thus, mining areas affected by long-term off-site release of As represent a serious concern for the environment and human health, particularly when used for grazing or agriculture. This situation is quite common as many former mine exploitations are located in areas where their main economic activities are livestock farming and agriculture. Arsenic is toxic to both plants and animals, with mammals being those most seriously affected. Arsenic toxicity depends on its chemical forms. Usually, inorganic species are more toxic than organic ones (Sharma and Sohn, 2009). Both As(III) and As(V) may cause similar toxicological effects, but the former is regarded as more toxic (KabataPendias and Mukherjee, 2007). Chronic exposure to inorganic As can lead humans to develop a variety of adverse health effects. Thus, its intake via food and/or water can cause skin changes or lesions such as hyperpigmentation, keratosis and ulceration, respiratory system problems, cardiovascular disease, nervous system alterations, hematological and immunological disorders, reproductive complications and cancer of different organs, namely, skin, lung and bladder, whereas its inhalation mainly impacts the respiratory and reproductive systems and the skin, including lung and skin cancer development as the primary adverse effect (Mandal and Suzuki, 2002; Basu et al., 2014). Ingestion of contaminated food and water is the main route of As intake into the organism (Kabata-Pendias and Mukherjee, 2007; Tarvainen et al., 2013; Sharma et al., 2014). The As content of many foods such as milk, beef, pork, poultry and cereals is mainly inorganic, typically reaching values of 65–75% of the total As content (Mandal and Suzuki, 2002). The relative content of inorganic As in vegetables has been shown to be much more variable, but levels close to 100% have been also reported (Mandal and Suzuki, 2002; Smith et al., 2006). In any case, pasture and agricultural lands impacted by former mine exploitations of arsenical deposits entail an important risk of inorganic As incorporation in the trophic chain. Many studies have been devoted to evaluate the As content and distribution in food-cultivated areas affected by past mining activities (e.g., Castro-Larragoitia et al., 1997; Liu et al., 2005, 2010; Williams et al., 2009; Álvarez-Ayuso et al., 2012; Li et al., 2010, 2014; Xue et al., 2017), revealing in numerous instances As accumulation in crops at levels that fail prescribed food standards. Nevertheless, studies on mine-impacted grazing lands are much more restricted (Li and Thornton, 1993; Abrahams and Thornton, 1994; Bruce et al., 2003; Tighe et al., 2005, 2013). Some of these studies have included trial experiments with animals to assess their exposure to As. Recently, the use of soil-plant-animal transfer models in mining areas has been suggested as an useful tool to establish the exposure risks of grazing animals, and finally humans, to toxic elements (MorenoJiménez et al., 2011; Rodrigues et al., 2012; Martínez-López et al., 2014; Simmler et al., 2016). Both soil ingestion and herbage intake are considered the possible entrance pathways of toxic elements into livestock grazing on lands affected by mining activities. In general, the ingestion of soil has been found the dominant via for the intake of As (Abrahams and Thornton, 1994; Bruce et al., 2003; Simmler et al., 2016), although levels around the half of the element intake has been also attributed to the consumption of grass (Rodrigues et al., 2012). This variability is highly dependent on the plant species, particularly on their specific characteristics of As uptake, transport and accumulation. Although usually plants that colonize As-polluted soils show an excluder behavior (Wang et al., 2002), the presence of plants able to accumulate intermediate As levels in their aerial parts is not so

2. Materials and methods 2.1. Study area The studied mining area is placed 2 km from the south of Barruecopardo village, in the north-west of the Salamanca province (Spain), where the Barruecopardo mine is situated. During the period 1912–1983, this mine exploited the largest tungsten deposit in Spain. This exploitation took place in two phases, in which two open pits of dimensions 800 m long × 12 m wide × 20 m deep and 500 m long × 40 m wide × 90 m deep were excavated, respectively (Fig. 1). The exploited deposit is a hydrothermal vein/stockwork hosted in granitic rocks. Veins are filled with quartz and an ore mineral assemblage of scheelite (CaWO4), the predominant tungsten-bearing mineral, wolframite ((Fe,Mn)WO4), pyrite (FeS2) and abundant arsenopyrite (FeAsS). Chalcopyrite (CuFeS2), molybdenite (MoS2) and cassiterite (SnO2) also occur locally (Arribas, 1979; Sanderson, 2008). Mining activities generated large amounts of wastes mainly composed of barren rocks, sulfide minerals (mostly arsenopyrite) and their weathering products. This mine emplacement is situated in an agricultural farming area, mainly devoted to the raising of cattle and sheep and, in a lesser extent, to cereal (rye, barley and wheat) cultivation. The climate of this area is Atlantic-continental, with long and cold winters, and short, warm and dry summers. Average seasonal temperatures vary between 5 and 8 °C in winters and 21 and 24 °C in summers. Most rainfall occurs from early autumn to mid-spring, with an annual precipitation of about 900 mm. Previous studies have been performed in this mining area seeking for new plant species for phytoremediation (Otones et al., 2011a). Sites affected by mining activities carried out throughout the second phase of mine exploitation and not used specifically as grazing lands were employed in such studies. Agrostis castellana (Boiss. & Reut.) and Scirpus holoschoenus L. were identified as suitable species to be used in As phytostabilisation strategies. 2.2. Sampling Soil and plant sampling was carried out in a grazing land located near a dump where mine wastes were accumulated during the first 229

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SAMPLING SITE X (100 m)

X (0 m)

Mine waste dump

50 m

250 m

BARRUECOPARDO MINING AREA Fig. 1. Barruecopardo mining area and location of sampling site.

established by Wenzel et al. (2001), available P was analyzed by the Olsen method (Olsen et al., 1954), and particle size distribution was determined by the pipette method (Gee and Bauder, 1986).

stage of mine exploitation (Fig. 1). This land is currently being used by local farmers for grazing cattle and sheep. Soil sampling was performed at 0, 10, 25, 50, 75 and 100 m from the mine dump. Profile soil samples in the depth up to 50 cm were taken using a 3-cm diameter core sampler. Ten profile soil samples were collected at each indicated sampling distance. Sampled cores were divided into depths of 0–10, 10–20, 20–30, 30–40 and 40–50 cm. Single composite samples corresponding to the different soil depths were produced for each of the mentioned sampling distances. A total of 30 soil samples were generated. Such soil samples were air-dried and sieved through a 2-mm mesh sieve prior to their characterization. Plant sampling involved the most characteristic pasture plant species grown in the studied grazing land, namely, Agrostis truncatula Parl., Holcus annus subsp. setiglumis (Boiss. & Reut.) M. Seq & Castrov. and Leontodon longirostris (Finch & P.D. Sell) Talavera. These plant species were collected from each soil-sampling site, this is at 0, 10, 25, 50, 75 and 100 m from the mine dump. Although grazing animals only consume their aerial parts, both roots and above-ground tissues were sampled in order to study the trace element translocation in these selected plant species. At least twenty plant specimens of each species were collected at the different sampling sites. Plant samples were separated into roots and aerial parts. A total of 36 plant samples were generated. These plant samples were firstly washed with tap water, then cleaned by means of an ultrasonic bath to remove soil dust particles, and finally rinsed with deionized water. Afterwards, cleaned plant samples were dried at 70 °C for 48 h, and finely ground for analysis.

2.4. Soil total and soluble concentrations of As and other trace elements Total and soluble concentrations of As were analyzed in all the collected soil samples. Additionally, other trace elements which could be present in the studied mining area (Cu and Pb) were also analyzed for their total and soluble concentrations. Total trace element concentrations were determined in triplicate by aqua regia digestion of finely ground soil samples, employing a Milestone Ethos Plus microwave oven operating at a temperature of 190 °C for 15 min. Trace elements were analyzed in the derived extracts by inductively coupled plasma-atomic emission spectrometry (ICP-AES), using a Varian 720-ES apparatus. The accuracy of the digestion and analytical methods was evaluated on the basis of standard reference materials (SRM 2709a and SRM 2711a); analytical errors < 10% were shown. Soluble trace element concentrations were determined in triplicate according to the European leaching standard test EN-12457-4 (2002). Extracts resulting from this process were analyzed for trace elements by ICP-AES. In the case of As, when its concentrations were below the ICP-AES detection limit (0.1 mg L−1) analyses were carried out by Electrothermal Atomic Absorption Spectrometry (ETAAS), employing a Varian Spectra AA-220 instrument equipped with a GTA 110 graphite atomizer (detection limit: 1 µg L−1). 2.5. Partitioning of As between soil fractions

2.3. Soil physicochemical characterization Soil samples collected at 0, 50 and 100 m from the mine dump underwent in triplicate the sequential extraction procedure developed by Wenzel et al. (2001) for As fractionation in soils. This procedure, composed of five extraction steps, was applied as described by Otones et al. (2011b). The extracts resulting from the successive extraction steps ((1) non-specifically adsorbed, (2) specifically adsorbed, (3) associated to amorphous/poorly-crystalline Fe and Al oxides, (4) associated to crystalline Fe and Al oxides, and (5) present in residual phases) were analyzed for As by ICP-AES or ETAAS, depending on the As concentration level.

Soil samples collected at 0, 50 and 100 m from the mine dump were characterized for their main physicochemical properties. Such properties were determined in triplicate as follows: pH was analyzed potentiometrically using the saturated paste method, organic matter (OM) was obtained by wet digestion following the Walkley-Black method (Walkley, 1947), the content of amorphous/poorly-crystalline Fe and Al oxides was determined by extraction with ammonium oxalate under darkness and that of Fe and Al crystalline oxides by extraction with ammonium oxalate/ascorbic acid according to the procedure 230

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2.6. Plant concentrations of As and other trace elements

3.2. Soil environmental characterization

Trace element (As, Cu and Pb) concentrations in plant samples were determined in triplicate by microwave digestion in a mixture of HNO3:H2O2 (8:2). This digestion process was performed in two steps, a first step at a temperature of 75 °C for 10 min, and a second one at 109 °C for 30 min. The derived extracts were analyzed for trace elements by ICP-AES. In the case of As, when its concentrations were below the ICP-AES detection limit analyses were carried out by ETAAS. The accuracy of the digestion and analytical methods was evaluated on the basis of the standard reference material SRM 1573a; analytical errors < 10% were yielded. A standard bioaccumulation factor (BF) (Silva Gonzaga et al., 2006) was determined for the soil total and soluble trace element concentrations (BFtotal and BFsoluble, respectively) as the ratio of plant trace element concentrations to soil total and soluble trace element concentrations, respectively. In both cases the trace element concentrations present in the uppermost soil layer were employed. The translocation factor (TF) was calculated as the ratio of trace element concentrations in plant aerial parts to those in plant roots. Arsenic speciation studies were also performed for plant roots in order to explain As translocation to plant aerial parts and its difference between plant species. The method of Zhang et al. (2002) was used with this aim. Accordingly, root samples were ultrasonically extracted twice with a methanol:water mixture (1:1) for 2 h at 60 °C. Resulting extracts were passed through As speciation cartridges (MetalSoft Center, NJ, USA), which retain arsenate (Meng et al., 2001). Extracts were analyzed for As by ETAAS.

3.2.1. Soil total and soluble concentrations of As and other trace elements The total and soluble trace element (As, Cu and Pb) concentrations in soils of the studied grazing land are reflected in Table 2. Total As concentrations found in soils of this area showed values comprised within the range 138–854 mg kg−1. Such concentrations importantly exceeded the critical As concentration range above which this element can cause toxicity (20–50 mg kg−1; Kabata-Pendias and Pendias, 1992). A good positive correlation was found between soil total As concentrations and soil total contents of Fe oxides (R = 0.813; p < 0.001). It is well known that these phases act as the major sink for As in soils (Bissen and Frimmel, 2003). Average As concentrations of 5.8 mg kg−1 have been reported for the uppermost layer (0–10 cm) of grazing land soils in Europe (Tarvainen et al., 2013). Much higher values were found in topsoils of the studied grazing area (138–379 mg kg−1). These concentrations showed a decreasing trend with the increasing distance from the mine dump. Thus a negative correlation was found between total As concentrations in the uppermost soil layer and distance from the mine dump (|R| = 0.910; p < 0.02). This is in agreement with the study performed by Xue et al. (2017) where the spatial distribution of As in a mining area was also found to be influenced by this anthropogenic activity. In any case, total As concentrations in topsoils of the studied grazing area greatly surpassed international soil guideline values for As in grazing land soils (50 mg kg−1; German Federal Soil Protection Act, 1998), suggesting the potential risk of As incorporation in the trophic chain from this pasture area. Total Cu and Pb concentrations found in soils of this area showed values of 23.6–101 and 20.0–64.3 mg kg−1, respectively. These concentrations were within their normal concentrations in soils (2–250 and 2–300 mg kg−1, respectively), and below or within their critical soil concentration ranges (60–125 and 100–400 mg kg−1, respectively). Soluble As concentrations in these soils ranged between 0.04 and 4.7 mg kg−1, showing, in general, a decreasing trend with depth, particularly in the close environs of mine dump. Such soluble concentrations represented ≤ 2% of the total As contents, as is usual in mine soils (Camm et al., 2004; Ongley et al., 2007; Pfeifer et al., 2007). However, even if the soil soluble As fraction showed quite low values, the highest soluble As concentrations, mostly found in the uppermost soil layers, were about two orders of magnitude higher than those found in uncontaminated soils (0.01 mg kg−1; Anawar et al., 2008). Such high soluble concentrations represent an important risk to the surrounding ecosystem. Usually, As concentrations in non-polluted waters exhibit values < 10 μg L−1, and often < 1 μg L−1 (Smedley and Kinniburgh,

3. Results and discussion 3.1. Soil physicochemical characterization The main soil physicochemical properties are given in Table 1. Soils of the studied grazing land exhibited an acid character, with pH values ranging from 3.8 to 5.0. Their OM content was moderate, showing in the uppermost soil layer values in the range 1.3–1.8%. Their available P was restricted (< 11 mg kg−1), presenting generally very low values (< 5 mg kg−1). The soil content of amorphous and crystalline Fe oxides reached considerable levels (0.18–0.70% and 0.18–1.4%, respectively), while that of amorphous and crystalline Al oxides was much more reduced (0.03–0.09% and 0.06–0.17%, respectively). Soils of this area had either heavy or medium textures, with the uppermost layers being the highest in clay content. Table 1 Soil physicochemical characteristics. Distance (m)

Depth (cm)

pH

OM (%)

Amorphous Fe2O3 (%)

Crystalline Fe2O3 (%)

Amorphous Al2O3 (%)

Crystalline Al2O3 (%)

Sand (%)

Silt (%)

Clay (%)

Available P (mg kg−1)

0

0–10 10–20 20–30 30–40 40–50

4.3 ± 0.1 3.9 ± 0.1 4.1 ± 0.1 3.8 ± 0.1 3.8 ± 0.1

1.8 ± 0.3 1.6 ± 0.3 2.0 ± 0.4 1.0 ± 0.1 0.98 ± 0.09

0.24 ± 0.03 0.28 ± 0.01 0.31 ± 0.08 0.70 ± 0.01 0.60 ± 0.06

0.18 ± 0.01 0.21 ± 0.01 0.20 ± 0.03 0.41 ± 0.01 0.41 ± 0.10

0.03 ± 0.00 0.05 ± 0.01 0.07 ± 0.02 0.09 ± 0.01 0.08 ± 0.01

0.06 ± 0.00 0.09 ± 0.01 0.11 ± 0.03 0.09 ± 0.03 0.10 ± 0.03

13.2 ± 1.6 23.7 ± 1.9 33.7 ± 2.4 35.3 ± 2.2 34.7 ± 2.2

19.3 ± 1.7 38.4 ± 2.5 51.1 ± 3.1 44.1 ± 3.0 39.8 ± 2.7

67.5 ± 2.0 37.9 ± 1.8 15.2 ± 1.3 20.6 ± 1.9 25.5 ± 2.2

4.6 ± 0.4 3.7 ± 0.5 4.7 ± 0.4 1.5 ± 0.3 0.77 ± 0.12

50

0–10 10–20 20–30 30–40 40–50

4.8 ± 0.1 5.0 ± 0.1 5.0 ± 0.1 4.9 ± 0.1 4.8 ± 0.1

1.6 ± 0.2 0.86 ± 0.08 0.65 ± 0.05 0.40 ± 0.05 0.20 ± 0.03

0.22 ± 0.01 0.20 ± 0.01 0.29 ± 0.06 0.31 ± 0.03 0.23 ± 0.02

0.57 ± 0.02 0.48 ± 0.05 0.77 ± 0.06 1.4 ± 0.1 1.3 ± 0.3

0.03 ± 0.00 0.03 ± 0.00 0.07 ± 0.04 0.08 ± 0.01 0.08 ± 0.01

0.08 ± 0.01 0.08 ± 0.01 0.10 ± 0.01 0.16 ± 0.01 0.15 ± 0.03

17.9 ± 2.0 15.7 ± 1.2 33.6 ± 1.8 56.5 ± 2.8 50.6 ± 3.0

14.6 ± 1.5 13.7 ± 1.3 18.0 ± 1.0 17.5 ± 1.1 17.9 ± 1.1

67.5 ± 2.9 70.6 ± 3.1 48.4 ± 2.4 26.0 ± 2.2 31.5 ± 1.8

0.91 ± 0.14 10.8 ± 1.0 0.49 ± 0.10 2.4 ± 0.3 n.d.

100

0–10 10–20 20–30 30–40 40–50

4.6 ± 0.1 4.7 ± 0.1 4.7 ± 0.1 4.7 ± 0.1 4.7 ± 0.1

1.3 ± 0.2 1.0 ± 0.3 0.96 ± 0.07 0.31 ± 0.05 0.22 ± 0.05

0.20 ± 0.02 0.29 ± 0.01 0.33 ± 0.02 0.22 ± 0.01 0.18 ± 0.02

0.36 ± 0.04 0.57 ± 0.01 1.2 ± 0.2 0.99 ± 0.20 0.90 ± 0.22

0.04 ± 0.01 0.05 ± 0.01 0.08 ± 0.01 0.07 ± 0.01 0.05 ± 0.01

0.09 ± 0.01 0.09 ± 0.01 0.15 ± 0.01 0.17 ± 0.01 0.14 ± 0.02

17.5 ± 1.4 4.6 ± 0.6 45.7 ± 3.1 47.9 ± 2.7 41.8 ± 2.9

13.8 ± 1.2 33.7 ± 1.9 13.2 ± 1.4 12.2 ± 1.2 11.8 ± 1.0

68.7 ± 2.6 61.7 ± 2.7 41.1 ± 2.0 39.9 ± 1.9 46.4 ± 1.7

1.8 ± 0.2 10.1 ± 0.9 n.d. n.d. n.d.

n.d.: not detectable.

231

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Table 2 Total and soluble As, Cu and Pb concentrations in soils. Distance (m)

Depth (cm)

Astotal (mg kg−1)

Cutotal (mg kg−1)

Pbtotal (mg kg−1)

Assoluble (mg kg−1)

Cusoluble (mg kg−1)

Pbsoluble (mg kg−1)

0

0–10 10–20 20–30 30–40 40–50

337 ± 63 153 ± 11 167 ± 31 411 ± 5 388 ± 28

60.0 ± 2.0 80.3 ± 3.6 101 ± 10 99.4 ± 1.9 92.8 ± 7.6

31.9 ± 2.5 48.5 ± 5.7 62.1 ± 0.8 64.3 ± 4.4 45.5 ± 1.7

1.61 ± 0.25 0.10 ± 0.05 0.06 ± 0.01 0.04 ± 0.00 0.07 ± 0.01

0.47 ± 0.01 0.85 ± 0.13 0.70 ± 0.27 1.02 ± 0.31 1.17 ± 0.08

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

10

0–10 10–20 20–30 30–40 40–50

379 ± 7 496 ± 9 268 ± 41 741 ± 7 782 ± 16

95.0 ± 1.6 78.8 ± 0.1 51.5 ± 4.4 99.4 ± 3.4 71.8 ± 9.3

30.4 ± 4.0 26.7 ± 2.7 32.8 ± 0.6 38.6 ± 0.9 37.7 ± 4.8

1.70 ± 0.01 0.62 ± 0.33 0.08 ± 0.01 0.06 ± 0.01 0.07 ± 0.01

0.55 ± 0.01 0.16 ± 0.05 0.23 ± 0.05 0.09 ± 0.04 < 0.05

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

25

0–10 10–20 20–30 30–40 40–50

365 ± 6 350 ± 13 389 ± 32 548 ± 29 518 ± 3

39.8 ± 0.8 31.1 ± 4.7 32.8 ± 3.3 48.0 ± 6.0 51.9 ± 0.4

27.2 ± 0.7 21.9 ± 2.2 35.6 ± 1.2 51.9 ± 5.9 42.7 ± 1.4

1.27 ± 0.14 1.58 ± 0.32 1.04 ± 0.04 0.52 ± 0.31 0.16 ± 0.05

0.40 ± 0.10 0.41 ± 0.03 0.23 ± 0.01 0.06 ± 0.02 < 0.05

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

50

0–10 10–20 20–30 30–40 40–50

293 ± 48 285 ± 40 430 ± 52 550 ± 4 854 ± 42

25.4 ± 1.3 23.6 ± 2.1 38.0 ± 4.4 55.1 ± 0.9 69.6 ± 2.1

24.7 ± 2.3 25.1 ± 2.4 38.2 ± 2.5 39.5 ± 0.1 45.1 ± 0.5

1.49 ± 0.22 1.97 ± 0.14 1.79 ± 0.09 0.76 ± 0.34 0.98 ± 0.06

0.31 ± 0.02 0.31 ± 0.02 0.25 ± 0.02 0.15 ± 0.08 0.44 ± 0.11

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

75

0–10 10–20 20–30 30–40 40–50

138 ± 37 256 ± 41 391 ± 54 272 ± 10 313 ± 31

24.5 ± 0.2 42.4 ± 3.6 53.9 ± 3.8 47.8 ± 3.6 54.7 ± 4.6

27.1 ± 4.4 38.8 ± 3.3 49.1 ± 5.0 23.9 ± 5.4 20.0 ± 2.7

1.98 ± 0.42 2.34 ± 0.29 1.39 ± 0.31 0.98 ± 0.02 3.28 ± 0.16

0.45 ± 0.12 0.38 ± 0.04 0.30 ± 0.13 0.30 ± 0.04 0.25 ± 0.07

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

100

0–10 10–20 20–30 30–40 40–50

157 ± 20 227 ± 14 457 ± 5 393 ± 28 379 ± 47

35.4 ± 1.9 48.3 ± 0.8 85.2 ± 0.3 80.4 ± 1.4 62.0 ± 7.3

22.4 ± 3.5 24.9 ± 0.4 32.1 ± 5.3 30.3 ± 0.3 31.3 ± 5.6

2.18 ± 0.08 4.56 ± 0.25 4.71 ± 0.21 0.30 ± 0.00 0.14 ± 0.06

0.61 ± 0.10 1.17 ± 0.12 1.37 ± 0.28 0.08 ± 0.02 0.06 ± 0.01

< 0.5 < 0.5 < 0.5 < 0.5 < 0.5

amorphous Al oxides and, to a smaller extent, amorphous/poorlycrystalline Fe oxides and crystalline Al oxides. These phases adsorb arsenite through the formation of outer-sphere surface complexes, being the only mechanism involved when adsorption takes place on amorphous Al oxides (Arai et al., 2000; Goldberg and Johnston, 2001; Voegelin and Hug, 2003; Wang and Mulligan, 2008). The As pool removed in the second stage, which corresponds to specifically adsorbed As, attained levels comprised between 3.9% and 19.5%. Different soil components can adsorb As tightly via an inner-sphere mechanism, such as Fe and Al oxides, non-crystalline aluminosilicates and, in a lesser extent, clay minerals (McBride, 1994). Apart from adsorption/desorption processes, As mobility in soils is also importantly controlled by co-precipitation with metal oxides. In this regard, the third and fourth extraction steps target As released after the dissolution of amorphous/ poorly-crystalline oxides of Fe and Al and crystalline oxides of Fe and Al, respectively. The corresponding As partitioning in such fractions reached values of 9.4–55.3% and 8.5–48.6%. According to the derived extraction levels, the general As partitioning between the different soil fractions followed the sequence: present in residual phases > associated to amorphous/poorly-crystalline oxides of Fe and Al > associated to crystalline oxides of Fe and Al > specifically adsorbed > non-specifically adsorbed. For soils next to the mine dump this sequence was slightly modified, being the As fraction associated to amorphous/ poorly-crystalline oxides of Fe and Al the prevailing phase. This agrees with the oxidation of arsenopyrite present in the mine dump to scorodite (FeAsO4·2H2O) and its further incongruent dissolution at pH ≥ 3, like those exhibited by soils. Such dissolution provokes the precipitation of Fe oxides and the release of As which is retained by the precipitated phases (Dove and Rimstidt, 1985; Nordstrom and Archer, 2003).

2002). These levels were importantly surpassed by As concentrations in leachates coming from the uppermost soil layer of the studied grazing land (127–218 μg L−1). Likewise, these leachate As concentrations were also above the recommended limit for As in irrigation water (100 μg L−1; Rowe and Abdel-Magid, 1995). Nonetheless, low groundwater As concentrations (mostly < 1 μg L−1) have been reported in this mining district (García-Sánchez and Alvarez-Ayuso, 2003). Aquifers in the study area are confined in granites and occur in fractured rocks with no external recharge from other aquifers (GarcíaSánchez and Alvarez-Ayuso, 2003). On the other hand, soluble As concentrations in soils have been reported to show a direct relationship with phytotoxicity (Woolson et al., 1971; Deuel and Swoboda, 1972). Thus, non-tolerant plant species can be injured at tissue As concentrations of 5–20 mg kg−1, with a likely 10% depression in yield even from As concentrations of 1 mg kg−1 in their tissues (McNicol and Beckett, 1985). Soluble Cu and Pb concentrations in these soils showed values of < 0.05–1.37 and < 0.5 mg kg−1, respectively, with all those in the uppermost soil layer being below the limit values recommended for soluble Cu and Pb concentrations in agricultural soils (0.7 and 1 mg kg−1, respectively; Ewers, 1991). 3.2.2. Partitioning of As between soil fractions The partitioning of As between the different soil fractions of the studied grazing land is indicated in Fig. 2. According to the results obtained from the applied sequential extraction, As was importantly partitioned in the residual fraction, with values varying between 11.5% and 61.0%. Conversely, low As levels (< 1%) were released in the first extraction stage, which targets labile or easily exchangeable As. The main soil components responsible for As adsorption in labile forms are

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0m

50 m

100%

100%

80%

80%

60%

60%

40%

40%

20%

20% 0%

% As

0% 0-10

10-20

20-30

30-40

0-10

40-50

10-20

20-30

30-40

40-50

100 m 100% 80% 60% 40% 20% 0% 0-10

10-20

20-30

30-40

40-50

Depth (cm) Associated to amorphous Fe and Al oxides Non-specifically adsorbed Specifically adsorbed Associated to crystalline Fe and Al oxides Present in residual phases Fig. 2. Partitioning of As between soil fractions.

varied according to the following sequence: Agrostis truncatula (5.5–32.9 mg kg−1) > Leontodon longirostris (7.3–16.9 mg kg−1) > Holcus annus (5.2–9.0 mg kg−1). The BFtotal for such tissues showed values comprised between 0.01 and 0.12, whereas values of 2.6–20.5 were reached by the related BFsoluble. The root As concentrations were in the range 14.9–98.4 mg kg−1, varying according to the following sequence: Agrostis truncatula (25.2–98.4 mg kg−1) > Holcus annus (16.7–80.0 mg kg−1) > Leontodon longirostris (14.9–31.1 mg kg−1). The BFtotal values for roots were comprised between 0.04 and 0.51 and the related BFsoluble values between 6.8 and 58.0. Bioaccumulation factors were higher for roots than for aerial parts, but differences were not very marked, being only about 2–4-fold higher. The TF values varied from 0.06 to 0.89. Arsenic translocation was produced to a greater extent by the species Leontodon longirostris, which showed TF values of 0.46–0.89, whereas those corresponding to Agrostis truncatula and Holcus annus were included in the ranges 0.06–0.71 and 0.11–0.32, respectively. Arsenic speciation in plant roots is shown in Fig. 3. All plant species contained predominantly arsenate in their roots, with minor arsenite levels, which accounted only for up to 34%, 25% and 22% of root total As contents in Holcus annus, Agrostis truncatula and Leontodon longirostris, respectively. Except for hyperaccumulators, usually plants show limited As translocation from roots to shoots. This behavior is generally explained by a rapid arsenate reduction to arsenite in roots and its further complexation with thiols (Zhao et al., 2009). In any case, redox conditions have been reported to affect As speciation in plant roots.

Changing environmental conditions more prone to cause As mobilization from soils include the incorporation of competitive anions (mostly phosphate) and the evolution towards reducing conditions under which amorphous/poorly-crystalline oxides of Fe are dissolved. Therefore, As released from the first three extraction stages is that more liable to be mobilized. Accordingly, in the studied grazing land the mobilizable As amounts would represent 16–72% of the total As contents. The highest risk of As mobilization (49–72%) was found in soils next to the mine dump. 3.3. Pasture plant species study The As, Cu and Pb concentrations in roots and above-ground tissues of pasture plant species collected in the studied grazing land are given in Table 3. The BF and TF values are also shown. The As concentrations found in the aerial parts of collected pasture species varied between 5.2 and 32.9 mg kg−1, exceeding the typical As concentration range in terrestrial plants (0.02–7 mg kg−1; Alloway, 1995). Likewise, such values surpassed the As contents (up to 6.4, 1.7 and 0.97 mg kg−1) reported in pasture grass species (Cynodon dactylon, Chloris gayana and a mixture of Holcus lanatus, Lolium perenne and Festuca rubra, respectively) grown in other studied mine-impacted grazing lands with diverse As pollution levels (up to 21.1, 1760 and 461 mg kg−1, respectively) (Li and Thornton, 1993; Bruce et al., 2003; Tighe et al., 2005). Arsenic accumulation in the above-ground tissues of studied plants

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Table 3 Arsenic, copper and lead concentrations in pasture plant tissues and bioaccumulation and translocation factors. Plant species

Agrostis truncatula Leontodon longirostris Holcus annus

Agrostis truncatula Leontodon longirostris Holcus annus

Agrostis truncatula Leontodon longirostris Holcus annus

Plant tissue

As (mg kg−1)

BFtotal

BFsoluble

TF

Aerial parts Roots Aerial parts Roots Aerial parts Roots

18.7 (5.5–32.9) 57.1 (25.2–98.4) 12.1 (7.3–16.9) 19.3 (14.9–31.1) 6.8 (5.2–9.0) 35.3 (16.7–80.0)

0.08 0.20 0.05 0.08 0.03 0.16

(0.01–0.12) (0.12–0.27) (0.02–0.08) (0.04–0.11) (0.01–0.06) (0.06–0.51)

11.6 (3.3–20.5) 35.4 (12.8–58.0) 7.2 (5.1–10.0) 11.8 (6.8–19.3) 4.0 (2.6–5.4) 20.1 (8.4–36.7)

0.41 (0.06–0.71)

Aerial parts Roots Aerial parts Roots Aerial parts Roots

Cu (mg kg−1) 11.2 (6.9–25.1) 23.4 (12.7–66.0) 15.6 (10.8–24.6) 29.6 (15.0–73.7) 7.8 (6.1–10.9) 28.4 (12.7–56.8)

BFtotal 0.26 (0.10–0.42) 0.53 (0.18–1.1) 0.38 (0.19–0.51) 0.67 (0.27–1.2) 0.19 (0.11–0.26) 0.63 (0.35–0.95)

BFsoluble 24.9 (13.2–54.0) 51.4 (23.1–142) 34.6 (21.9–52.7) 64.2 (38.3–158) 17.1 (11.9–19.7) 60.0 (28.5–122)

BFtotal < 0.63 < 0.63 < 0.63 < 0.63 < 0.63 < 0.63

BFsoluble

Aerial parts Roots Aerial parts Roots Aerial parts Roots

Pb (mg kg−1) < 20 < 20 < 20 < 20 < 20 < 20

0.65 (0.46–0.89) 0.23 (0.11–0.32)

TF 0.54 (0.38–0.63) 0.62 (0.33–0.86) 0.34 (0.15–0.53)

TF

- < 0.89 - < 0.89 - < 0.89 - < 0.89 - < 0.89 - < 0.89

According to the Directive 2002/32/EC on undesirable substances in animal feed (2002), the As concentration in the above-ground tissues of the different pasture species of this grazing land exceeded the maximum As content (2 mg kg−1 fresh weight (FW), equivalent to 2.3 mg kg−1 dry weight (DW), considering a moisture content of 12%) allowed in most animal feed for avoiding its transfer and further accumulation in the higher trophic levels of food chain. Therefore, the use of this land for grazing should be considered unsafe given the potential for As accumulation in grazing animals trough plant ingestion. The risk posed by the studied pasture species follows the order: Agrostis truncatula > Leontodon longirostris > Holcus annus, as derived from As concentrations in their above-ground tissues. Lead is also considered in the Directive 2002/32/EC on undesirable substances in animal feed (2002), establishing a maximum Pb content of 40 mg kg−1 FW in green fodder (equivalent to 45.5 mg kg−1 DW, assuming a moisture content of 12%). Lead concentrations in the different pasture species of the studied grazing land were well below this fixed limit.

Thus, root arsenite concentrations of different rice genotypes have been found lower when plants were grown under oxic conditions than when grown in anoxic treatments (Wu et al., 2017). The reduced arsenite fractions in plant roots of the studied pasture area could explain the relatively high TF values exhibited by these plant species. In this regard, the sequence of species showing increasing TF values (Holcus annus < Agrostis truncatula < Leontodon longirostris) agreed with that exhibiting decreasing arsenite fractions in their roots (Holcus annus > Agrostis truncatula > Leontodon longirostris). The Cu concentrations found in the aerial parts of collected pasture species varied between 6.1 and 25.1 mg kg−1, being within or close to the typical Cu concentration range in terrestrial plants (5–20 mg kg−1; Alloway, 1995). Copper accumulation levels in the above-ground tissues of studied plants varied according to the sequence: Agrostis truncatula (6.9–25.1 mg kg−1) > Leontodon longirostris (10.8–24.6 mg kg−1) > Holcus annus (6.1–10.9 mg kg−1). The BFtotal for such tissues showed values comprised between 0.10 and 0.51, whereas values of 11.9–54.0 were reached by the related BFsoluble. The root Cu concentrations were in the range 12.7–73.7 mg kg−1, showing quite small differences between plant species. The BFtotal values for roots were comprised between 0.18 and 1.2 and the related BFsoluble values between 23.1 and 158. Bioaccumulation factors for roots were about twice those for aerial parts. The TF values varied from 0.15 to 0.86. Like As, Cu translocation was produced to a greater extent by the species Leontodon longirostris. The Pb concentrations in both roots and aerial parts of collected pasture species showed values < 20 mg kg−1. Lead concentrations in aerial tissues were within the typical Pb concentration range in terrestrial plants (0.2–20 mg kg−1; Alloway, 1995), as could be expected due to the relatively low soil total and soluble Pb concentrations in this mining area and to the usual low transference from roots to shoots shown by plants for this trace element (Adriano, 1986). Previous studies have reported plant uptake of multiple trace elements in areas impacted by mining activities (e.g., Freitas et al., 2004; Chang et al., 2005; Antosiewicz et al., 2008; Marques et al., 2009). Some of these studies have identified plants with great potential to be used in phytoremediation strategies, and others have pointed out the risks associated with the high accumulation levels of either single or various trace elements in plant shoots.

3.4. Exposure of grazing animals to As Both soil ingestion and pasture intake are considered the possible entrance pathways of trace elements into grazing animals. The total daily intake of As by grazing animals and the relative contribution of each pathway can be established according to the equation described by Smith et al. (2009). It is as follows:

DItotal = DIsoil + DIpasture = Isoil x [As]soil + Ipasture x [As]pasture

(1)

where DItotal, DIsoil and DIpasture are the total daily intake of As, the As daily intake via soil and the As daily intake via pasture (mg d−1), respectively, Isoil and Ipasture are the intake of soil and pasture (kg DW d−1), respectively, and [As]soil and [As]pasture are the As concentrations in soil and pasture (mg kg−1 DW), respectively. The Isoil and Ipasture values used to calculate the DItotal, DIsoil and DIpasture of animals grazing in the studied land (cow and sheep) are those given by McKone and Ryan (1989) and de Vries et al. (2007) (cow Isoil and Ipasture = 0.41 and 16.9 kg DW d−1, respectively, and sheep Isoil and Ipasture = 0.10 and 2.5 kg DW d−1, respectively). The calculated DItotal, DIsoil and DIpasture of As and the relative

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Agrostis truncatula

important intake of As trough pasture is in agreement with the results from the study of Rodrigues et al. (2012) where pasture grown in soils polluted with As also contributed significantly to the total As uptake (50% (cow) and 38% (sheep)). In order to assess the risks that may entail the derived DItotal values, these were compared with the acceptable daily intake (ADI) values for As regarding food safety and animal health, using the minimal ADI values given by Rodrigues et al. (2012) (cow and sheep ADIfood safety = 500 and 182 mg d−1, respectively, and cow ADIanimal health = 3502 mg d−1). Such values were calculated taken into consideration the limit concentration recommended for food safety and animal health (López Alonso et al., 2000; Nriagu et al., 2009). Particularly, for food safety an As limit concentration of 2 mg kg−1 FW is recommended for cow kidney and liver (Nriagu et al., 2009) (the limit for bovine organs was assumed to be equal to this value), and for animal health an As limit concentration of 14 mg kg−1 FW is recommended for cow kidney and liver (López Alonso et al., 2000). The normal As concentrations in kidney, liver and meat/muscle of cattle (< 0.015–0.068, < 0.015–0.050 and 0.004 - < 0.02 mg kg−1 FW, respectively; Miranda et al., 2003) and sheep (< 0.001–0.044, < 0.001–0.025 and < 0.001–0.004 mg kg−1 FW, respectively; Vos et al., 1988; Uneyama et al., 2007) show values well below the indicated limits for food safety and animal health. According to the given ADI, the cow and sheep ADI as regards food safety and the cow ADI as regards animal health were not exceeded by the average DItotal values. Nonetheless, the cow ADI as regards food safety was surpassed in some locations of the study area when the species Agrostis truncatula was considered as the only pasture feed. Therefore, grazing in areas where this species was the dominant pasture herbage should be avoided in order to assure safe food levels.

100%

80% 60% 40% 20% 0% 0

10

25

50

75

100

Holcus annus 100%

80% 60% 40% 20%

4. Conclusions

0% 0

10

25

50

75

The environmental characterization of a pasture area impacted by the former mine exploitation of an arsenical tungsten deposit indicated the presence of highly increased As concentrations in soils (138–854 mg kg−1), greatly surpassing international soil guideline values for As in grazing land soils. Also the soil soluble As concentrations reached high values (up to 4.7 mg kg−1), pointing out an important risk of As transference to other environmental compartments. Moreover, under the presence of competing anions and/or the evolution towards reducing conditions the mobilizable As amounts could attain values of 16–72% of the total As content. Pasture species grown in this field accumulated important As levels in their aerial parts (5.2–32.9 mg kg−1), following the sequence: Agrostis truncatula > Leontodon longirostris > Holcus annus. According to international legislation, such concentrations exceeded the maximum As content allowed in most animal feed. Therefore, the use of this kind of mine-impacted lands for grazing could entail a high risk of As transfer and accumulation in the higher trophic levels of food chain. Thus, the consumption of pasture was found an important pathway for the intake of As by grazing animals, attaining levels similar or even higher than those corresponding to soil ingestion. The increasing TF values in plant species corresponded with the decreasing arsenite fractions in their roots, suggesting the importance of As speciation in roots in As translocation. The cow acceptable daily intake (ADI) of As as regards food safety was exceeded in some locations of the study area when the species Agrostis truncatula was considered as the only pasture feed. Restrictions in the grazing use of lands with considerable As contents where this plant was the predominant pasture species should be established in order to preserve food quality. Hence, the identification of either highly As-polluted areas or pasture plants with capacity to uptake and translocate As is necessary in order to establish the suitable measures to prevent the As incorporation into the food chain from pasture areas.

100

Leontodon longirostris 100%

80% 60% 40% 20% 0% 0

10

25

50

75

100

Distance (m) As(III)

As(V)

Fig. 3. Arsenic speciation in pasture plant roots.

contribution of soil and pasture to the DItotal are given in Table 4. Assuming an average mixture of the three plant species as a good surrogate of pasture feed, the DItotal showed average values of 326 and 59 mg d−1, for cow and sheep, respectively, with maximum levels reaching values up to 695 and 116 mg d−1, respectively. The contribution of As intake via pasture consumption to the total As intake showed medium values of 62% (cow) and 51% (sheep). The relatively

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Table 4 Calculated daily intake values of As and relative contribution of soil and pasture to the total daily intake of As. Grazing animal

Plant species

DIsoil (mg d−1)

DIpasture (mg d−1)

DItotal (mg d−1)

Contributionsoil (%)

Contributionpasture (%)

Cow

Agrostis truncatula Leontodon longirostris Holcus annus Mixture of plant species

114 114 114 114

317 205 114 212

431 319 228 326

29 35 49 38

(17–62) (23–55) (30–63) (17–63)

71 65 51 62

(38–83) (45–77) (37–70) (37–83)

Sheep

Agrostis truncatula Leontodon longirostris Holcus annus Mixture of plant species

28 28 28 28

39 47 61 49

(25–73) (33–67) (41–74) (25–74)

61 53 39 51

(27–75) (33–67) (26–59) (26–75)

(57–155) (57–155) (57–155) (57–155)

(14–38) (14–38) (14–38) (14–38)

47 30 17 31

(94–556) (124–286) (88–152) (88–556)

(14–82) (18–42) (13–22) (13–82)

75 58 45 59

(249–695) (246–441) (144–286) (144–695)

(47–116) (42–80) (27–56) (27–116)

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