Environmental Impact of Abandoned Mine Waste: a review

2.4. Overbank Stream Sediments

Overbank river sediments also show a marked dilution with distance from the pollution source, as reported by Mascaro et al. (2001). Sediments deposited along rivers that drain mine areas (Figure 5) are often highly polluted due to both geochemical background and also to the industrial mining practices (Gonzales-Fernandez et al., 2011). Mobilization of metals is primarily controlled by pH and change in redox conditions between oxic waters and anoxic sediments, that may have profound influence on metal bioavailability, including metal complexation of organic and inorganic ligants (Aleksander-Kwaterczak and Helios-Rybicka, 2009). Changes in redox conditions may also trigger the transfer of toxic elements from the particulate phase to the solution. This occurs mainly during summer, when the increase in temperature favours the development of anoxic or suboxic conditions in sediments, and boosts bacterial activity. These conditions favour the reduction of oxide phases and the mobilization of associated metals. Changes in other parameters such as an increase in temperature and/or pH also favour metalloid desorption in AMD-affected water (Casiot et al., 2009). Sivry et al. (2010) report in floodbank soils higher enrichment factors relative to France average soil metal content as far as 1km downstream of mine wastes. The water in the proximity of the contamination source has acid pH values, and high contents of sulphates and metals, in particular Cu, Zn, Mn, Fe. The waters collected both upstream and downstream are neutral with lower metal contents. The range of water pH and metal contents are not ascribed to different possible pollution sources, but to a combined action of dilution with unpolluted water moving downstream of mine wastes, and of buffering by carbonate rocks that outcrop nearby (Mascaro et al., 2001). Arsenic release from river sediments downstream of a gold mining district, instead, is greatly influenced by elevated pH (Rubinos et al., 2010). A correlation between pollutant transport and rainfall was also observed in other small basins affected by mining activities (Sanden et al., 1997). High dissolved concentrations of PHEs (SO₄^2-, Fe, Zn, Pb, As and REE) were found also in surface waters up to 1500m downstream from a mine site in Cuba (Romero et al., 2010).

The mineralogy of overbank sediments is mainly composed of quartz, feldspar, calcite and phyllites (i.e. clay minerals) as principal phases; the largest quantities of pyrite, chalcopyrite, galena, sphalerite, Fe oxides occur generally close to the waste and decrease downstream (Mascaro et al., 2001a). The ochreous muds frequently occurring at these sites consist of iron oxhydroxides (ferrihydrite) and quartz mechanically transported as suspended matter during flooding episodes; it is likely these muds to be responsible for the high metal content found in bottom sediments and in soils along the river overbank. As a general rule, bulk chemistry of sediments seems to be influenced by proximity to the mine waste; indeed, as far as 5 km downstream, sediments maintain relatively high contents of metals, in particular Zn.

Figure 5. Overbank sediments along the Imperina creek (Belluno, Northern Italy). (Photo Bini).

n2.5. Soil Contamination

Aquatic ecosystems are polluted by drainage from old mines, and erosion of mine waste or tailings still contribute to river sediments. The main impact of mobilisation of metal-rich materials from mine waste, however, is on the terrestrial ecosystems (Davies et al., 1983). One of the environmental compartments particularly sensitive to chemical contamination is soil. During mining operations, large amounts of waste (up to millions of cubic metres), dumps. heaps, tailings, metal-enriched and frequently strongly acidic (pH<3) waters, have been discharged in the surrounding land, determining degradation and contamination of soils. Wind blow is also a mechanism whereby toxic tailings can be transported to neighbouring agricultural land: Davies and White (1981) reported that most of the <2mm fraction of the spoil material in Wales was of sufficiently small diameter to move by deflation caused primarily by winds, and movement of spoils could be detected as far as 1800 m downvalley.

Figure 6. A deeply weathered mine dump in Sulcis, Sardinia (Photo Bini).

Mining operations affect relatively small areas. Actually, tailings and waste rock deposits close to the mining area are the main source of soil and water pollution (Salomons, 1995; Krzaklewski et al., 2004; Moreno-Jimenez et al., 2009). After extraction of economic metal ores, mine spoils resulting from mining works were dumped in close proximity to the mines and constitute a waste area on the modern landscape (Figure 6). The original surface soil was unevenly buried under mine tailings, so that natural processes of soil evolution were hindered. The waste surface remained uncovered for a long time, until weathering, revegetation and pedogenic processes enhanced soil formation in the mine spoil.

The rate of pedogenesis and the degree of soil evolution depend on several factors: the nature of the parent material, the residence time of parent material within the zone of active soil formation, the climate conditions, the soil hydrology (Moody and Graham, 1995). Materials derived from metal mines contain up to several weight percent of sulphidic minerals (Benvenuti et al., 1999) which, depending on the local hydrology, pH and redox status, upon oxidation and leaching, can generate strong acid conditions toxic to soils and plants, producing significant environmental impacts in the whole area (Benvenuti et al., 1999).

Figure 7. Mine waste with bare vegetation at the Temperino mining site, Campiglia Marittima, Tuscany. (Photo Bini).

After reaching the soil, metals are mainly accumulated in the upper organic and organic-mineral horizons. Mine soils are generally shallow and/or infertile soils which often are unsuitable for vegetation (Roberts et al., 1988). Coarse fragments form >70 % of the soil volume, and rooting is concentrated on coarse fragments faces. High coarse fragment contents reduce water availability and, therefore, soil evolution is very limited. According to Roberts et al. (1988), morphologically distinct A horizons, with weak granular structure, form in 5 years, but subsurface C horizons are undifferentiated; formation of cambic-like B horizons, with well-expressed blocky structure, but too shallow (<25cm) to meet cambic criteria, in 50 years-old mine soils, is reported by Schafer et al. (1980) in Montana, USA. Néel et al. (2003) found that the low rate of soil development (0.25 - 0.70 cm year^-1) from mine spoils in France could be related to inherited factors of parent waste materials such as the initial sulphide content, porosity, water content, texture, pH and redox conditions. The 35 years-old mine soils showed all the features of an immature A-C sequence, with a thin solum (<25 cm), little organic matter accumulation, a sandy-skeletal texture, acidic reaction, high metal contents (As 0.1-6%; Pb 0.2-2%; Sb 0.02-0.1%; Cu and Zn 30-200 ppm). All these properties change gradually with the distance from the waste discharge area.

Abandoned mine soils contain excessive contents of heavy metals, as it occurs in serpentine soils (Jenny, 1980; Raous et al., 2010). They have coarse grain size, poor moisture retention properties and lack of major nutrients. Owing to their high infertility, the abandoned mine soils are often bare of vegetation (Figure 7), and their steep sides make them unstable: yet, lacking vegetative cover renders mine tailings very prone to mobilization. Nevertheless, all plants take up metals to varying degrees from the substrates in which they are rooted. Metal concentrations in different plant tissues depend both on intrinsic (genetic) and extrinsic (environmental) factors, and vary greatly from plant to plant, and for different metals. The plants which colonize these soils are usually metal tolerant ecotypes, accumulator or hyperaccumulator plants (Baker, 1981), and their metabolic equilibrium is not altered by increased metal uptake (Adriano, 2001). Indeed, Bradshaw and Chadwick (1980) and, more recently, Chaney et al. (1995) have described how tolerant ecotypes may be used to revegetate metal contaminated soils. Yet, revegetation has been carried out successfully in temperate climate (see f.i. Madejon et al, 2002; Moreno-Jimenez et al., 2009). However, although visible toxic effects rarely extend beyond a few meters from the waste material, metals may be adsorbed by plants and could represent a potential contamination way of the food chain.