Chapter 1 - Home page | Archivio Istituzionale della Ricerca · Web viewAn absorption sequence Zn>Co>Cu>Ni>Fe>Cr, consistent with leaching tests, was found by Dinelli and Lombini (1996)

ENVIRONMENTAL IMPACT OF ABANDONED MINE WASTE:

A REVIEW

ENVIRONMENTAL IMPACT OF ABANDONED MINE WASTE:

A REVIEW

CLAUDIO BINIAUTHOR

Nova Science Publishers, Inc.New York

Copyright © 2011 by Nova Science Publishers, Inc.

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CONTENTS

PREFACE

ACKNOWLEDGEMENTS

The Author is indebted with collegues and cooperators that provided data for this review, and helped in preparing the final draft of the paper. Particular thanks are due to Dr. Mohammad Whasha, who revised the English form; Dr. Diana Zilioli and Dr. Silvia Fontana assisted in the field survey and provided laboratory analyses.

Prof. Jaume Bech, Chair of Soil Science, University of Barcelona (Spain), is warmly acknowledged for critical review and suggestions that contributed to improve an early draft of the paper.

Chapter 1

INTRODUCTION

Since the dawn of civilization and for long time, until the last decades of past century, mining activity, especially that concerning base (Cu, Fe, Pb, Zn) and precious metals (Au, Ag), as reported by George Bauer, (known as Agricola), in his book De re metallica (1556), represented a resource for human population, owing to its importance in many fields of interest: economic, cultural, technological (Figure 1).

By the second half of last century, however, mining activity, almost in European countries, declined until final closure, in the face of developing countries, owing to decreasing mineral resources, and to metal price drop. Since then, arose the problem of visible reminders and invisible inheritance of mine working (Davies, 1987), with reference to different aspects:

Environmental: soil contamination by metals, soil and water acidification; damage to vegetation;

Geomorphologic: landscape modification, geological hazard (erosion, flooding, landslides);

Sanitary: risk for human health (inhalation, ingestion, contact); Casual/professional diseases: intoxication, lead poisoning,

mercurialism.

Quite recently, however, abandoned mine sites have been discovered to constitute a chance, giving the opportunity to open Mine Parks and

Museums; Archaeological Parks; protected natural areas, didactic-recreational itineraries, trekking areas, and other activities in open air. Yet, mine sites are actually natural scientific laboratories, where to explore natural processes involving rock-forming minerals, their transformation into soil-forming minerals, their interaction with organic matter, and fluxes from soil to plants. Furthermore, mine sites investigations have been addressed to soil remediation and environmental restoration, for example with application of phytoremediation technologies (Bini, 2009).

Figure 1. The front of the treatise De re metallica by G. Agricola (1556).

Introduction

More recently, the European Mine Waste Directive (EC, 2006) has introduced new requirements for mine waste management, including that resulting from historical mining (Palumbo-Roe et al, 2009). The challenge in implementing the European Directive is to develop a pan-European risk-based inventory of abandoned mines, in order to select sites for remediation based on a common set of criteria. The characterisation of the mine waste and its transformations in the short and long term, forms the basis for a risk-based classification of abandoned mine sites (Servida et al., 2009).

In this paper, the effects of former mine activities, and the related environmental problems, with particular reference to Italy, are discussed, with the ultimate goal of investigating the fate of potentially toxic elements in the environment, and their impact on the conterminous land.

1.1. RESOURCE

Mineral exploitation, smelting and recovery of useful and/or precious metals in several countries of Europe dates back to VII century B.C. (Etruscan times) or even before (Thornton, 1996). After a large diffusion of Fe, Cu, Au, Ag, Sn, Pb mining during the Roman expansion in Europe and Britain, ore exploitation virtually ceased during the Middle Age (5 th to 11th centuries), and became economically important again after the 15 th

century, when there was an increasing demand for silver for coinage, and lead for armaments. (Davies, 1987). Afterwards, alternate fortunes accompanied mine works, particularly during the Industrial Revolution and until the first decades of 20th century, when mining activity in the Old Continent ceased and most mines were abandoned, for both exhaustion of metal veins, price drop and major sensitivity of people to human and environmental health. Silver and mercury, for instance, have been used since early historic times, as reported by Bargagli (1995) and Forel et al. (2010). Silver exploitation in the Vosges Mountains is attested since the 10th century in the Val d’Argent (NE France), where up to 600 mines have been accounted for at least 3000 miners (Forel et al., 2010). Cinnabar exploitation in the Mediterranean basin (Spain, Italy, Croatia, Turkey,

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Claudio Bini

Tunisia) began in Etruscan times, was expanded by Romans and dominated the world mercury production for long time (Bargagli, 1995). Mercury production, as well as that of other metals such as copper, lead, silver, zinc, etc. depended on the market price, which reached 700US dollars/kg during early ‘70s (Gemici et al., 2009). The gradual decline in the demand, caused by the increasing environmental concerns of Hg, resulted in lowered price, drastic reduction in mining, and the final closing of many Hg mines until the early ‘80s.

Metals are indissolubly linked to the progress of mankind, having greatly contributed to the evolution of civilization, from the stone age (Neolithic period, 6.000 BC), through copper, bronze, iron and “gold age” (the California gold rush), to present time. Exploitation of metals such as Cu, Au and Ag, for example, is among the most long lasting mining operations, since their recovery started with the chalcolithic age (copper-bronze age), between 5500 and 3000 BC (Dill, 2009).

Combining archaeological and geological investigations, numerous studies have focused on ancient settlements, artifacts and archaeometallurgical slags found at different sites, shading some light on the techniques applied for the recovery of pure metals (Cu, Au, Ag, Pb, Sb, Sn and Fe) from the various raw materials (Francovich, 1985; Stiles et al., 1995; Mascaro et al., 1995; Heimann et al., 1998; Manasse et al., 2001; Manasse and Mellini, 2002; Costagliola et al., 2008; Dill, 2009).

Table 1. Mine production of heavy metals

Element Mine production 1990antimony 55arsenic 45cadmium 19chromium 6800copper 8110lead 3100mercury 6.8

4

Introduction

nickel 778zinc 6040

Data is in metric tonnes x103 / year.Modified after McGrath, 1995.

Metals have been, and are still, mined in the majority of the countries of the world, and primary production of many metals continues to rise (Thornton, 1996). In 1950, the production of Pb was 1.7 million tonnes (Table 1), and that of Cu 2.8Mt, Cr 2.2Mt, Zn 1.9Mt, Ni 0.14Mt; in 1995, Pb production was up to 3.3 million tonnes; Cu 9.4Mt; Cr 12.8Mt; Zn 7.1Mt; Ni 0.9Mt (Thornton, 1996). The main reason for this interest towards metals is, obviously, related to their large utilization at worldwide level.

Nowadays, heavy metals are vital components of modern technology, being utilized in many industrial and agricultural activities (electronic, galvanic, metallurgy, varnish, tannery, wood preservation, fertilizers, pesticides, etc.) (Davies, 1987; Adriano, 2001). The metal over-utilization at worldwide level is responsible for serious threats to the environment, with potential risk for human health. Besides the occasional lead poisoning recorded during Roman domination ( Nriagu, 1983; Stiles et al., 1995), the first signals of threats appeared on agricultural land contaminated with heavy metals. In the middle of the 19th century, farmers living close to lead-silver mine areas in England complained that mine waste was deposited on fields by river floods, contaminating their land (Davies, 1980). It has been calculated that approximately 35% of the mineral waste discharged on the land was released to the environment. Overall, it can be estimated that for every ton of silver-free lead which was produced, as much as 2tons may have been lost to the environment (Davies, 1987). Similarly, Helios-Rybicka (1996) reported that approximately 700M tonnes for year of mineral commodities have been exploited in Poland, strongly influencing the hydrological system. Geomechanical processes (subsidence, slumps, landslides, erosion) led to the complete destruction of soils and irreversible changes of the landscape.

From that time, repeated threats to the environment have been recorded in current literature, (see f.i. Cappuyns et al., 2006; Palumbo-Roe et al., 2009), suggesting heavy metals to represent a concrete environmental concern.

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1.2. PROBLEM

Rocks and ore deposits are composed of a pool of chemical constituents. The major elements (Si, Al, Ca, Mg, etc.) are invariably accompanied by minor (Fe, Mn, Ti, P, etc.) or trace amounts of other metals. Among these, heavy metals can be defined (Adriano, 2001) as those having a metallic density >5 gcm-3 (e.g. Cu, Co, Ni, Pb, Zn, etc.). Other important constituents, particularly utilized in modern industrial activities, are antimony, arsenic, bismuth, cadmium, chromium, germanium, selenium, tallium, etc. In addition, many metals are essential for life functions. Chief concern focuses on Cu and Zn, which are essential micronutrients but may be harmful when present in large concentrations, and on Cd, Hg and Pb, which have no known beneficial metabolic role but are known toxins (Kabata-Pendias and Pendias, 2001; Ghorbel et al., 2010). The important point is that many of these metals are also potential contaminants to the environment, and constitute a potential risk to vegetation and human health, when their concentrations are above a certain threshold (Davies, 1987; Kabata-Pendias, 2004). Yet, these metals are ordinarily present in rocks, sediments and soils, but locally may become concentrated in rocks as ore bodies and generally dispersed in the environment through pollution as a consequence of mining the ores. (Davies, 1987; Alloway, 1995).

Mining is only one of the pathways by which metals enter the environment. Mining itself affects relatively small areas, and this could not pose problems. The environmental problem arises when ores are mined, milled and smelted, and a certain amount of metals is released in the surrounding areas and to waterways. Depending on the nature of the waste rock and tailings deposits, a wide dispersion of the metals both in solution and in particulate form is possible (Sivri et al., 2010). Erosion of waste rock deposits or the direct discharge of tailings in rivers results in the introduction of metals in particulate form into aquatic ecosystems (Helios-Rybicka, 1996; Cidu et al., 2009). Smelting of ore deposits results in the release of metals to the atmosphere (Mihalik et al., 2011); when metals

6

Introduction

have been released through the atmosphere, they end up as diffuse pollutants in soils and sediments. (Nriagu, 1990; Salomons, 1995).

Figure 2. Geological and hydrological hazard determined by mine dumps in the Metalliferous Hills district (Southern Tuscany, Italy). (Photo Bini).

A second environmental (geomorphologic) concern is connected to mining operations. Excavation of ore bodies brings out important landscape modifications; earth movements, dam building and impoundment construction, may create severe geological hazard (Figure 2). Erosion processes may concur to convey waste in rivers nearby; landslides may be activated in loose material dumps, with relevant risk for population living in the conterminous land; surface hydrology and hydrological processes may be strongly modified, and constitute a further concern.

Mining areas are frequently constituted of highly tectonized and fractured rocks and detrital fragments (Tanelli, 1985), easy erodible by runoff and percolating water. The causes of accelerated surface erosion are related to both geological (tectonic structure, lithology), morphologic and climatic conditions (steep slopes, rainy events distribution, temperature

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Claudio Bini

gradients); vegetation cover may have great influence on attenuating, or enhancing, erosion phenomena, when land cover is scarce or lacking, as it happens frequently with metal-contaminated sites. A correlation between tectonic structure, minerogenesis and surface processes has been recorded (Lattanzi et al., 1994; Benvenuti et al, 1999; Mascaro et al., 2000) at nearly every mine site.

Access to exploitation areas is allowed through earth movements such as opening new tracks and new excavations, and this may enhance earth surface processes, landslide formation, hydrological regime alteration, contaminant dispersion (Helios-Rybicka, 1996). The hydraulic characteristics of mineral bodies (e.g. coarse grain size, permeability, hydraulic conductivity), in turn, are responsible for water percolation and circulation in the subsoil, where contaminants are convoyed to groundwater (Cidu et al., 2009). Moreover, apart from geomechanical processes (subsidence, slumps, landslides, erosion) which may lead to the complete destruction of soils and irreversible changes in the landscape, the drainage of open pits influences hydrological systems (Helios-Rybicka, 1996).

A third set of problems, of sanitary type, may occur with workers involved in mining operations. Today, heavy metals are considered as hazardous substances which can induce environmental threats and a risk to human health (Ghorbel et al., 2010). Health risk assessment depends upon complex interactions of several parameters such as waste mineralogy, exposure, climate conditions, contamination transfer, contact, ingestion or inhalation of metals, time elapsed since mine closure.

Yet, if absorbed in sufficiently high amounts, heavy metals can be toxic and even lethal. The toxicity of heavy metals to humans is well documented by several outbreaks of massive poisoning and epidemiological studies (Steinnes, 2009). In particular Ag, As, Be, Cd, Ce, Ge, Hg, Pb, Tl are examples of potentially harmful elements (PHEs) that have no proven essential functions, and are known to have adverse physiological effects at relatively low concentrations (Abrahams, 2002).

Examples of toxicity by heavy metals are known since the Antiquity (Nriagu, 1983). For instance, one of the supposed causes for the Roman Empire drop is the increasing lead toxicity from Pb-bearing potteries and wine containers, as it was found in Roman findings and bones (Stiles et al., 1995). Lead (plumbism) and Hg (mercurialism) poisoning cases were

8

Introduction

frequently recorded in workers employed in mining industry and even in hat factories in Tuscany (Dall’Aglio et al., 1966). At present, diseases and toxicity related to microelement contamination (Cr, Cu, Ni, Pb, Tl, Zn,) of air, water and soil from human activities are well established (Thornton, 1993; Abrahams, 2002). For example, the most notable cause of Tl poisoning occurred adjacent to a cement works in Germany (Abrahams, 2002).

The history of lead and its use by man dates back to almost 9.000 years. The toxic nature of lead compounds was well understood, and there was a variety of local names for lead poisoning: plumbism, saturnism, potter’s rot, painters’colic, lead palsy (Davies, 1987). The main target for health hazard by lead are the hematopoietic, the nervous and the cardio-vascular systems (Bernard, 1995). Early scientific investigation of river pollution demonstrated that dissolved lead caused the formation of mucus on the gills of fish; stickleback (Gasterosteus aculeatus L.) was the species most sensitive to lead pollution (Davies, 1987).

Cadmium, in contrast, is typically a heavy metal of the 20 th century, since over 60% of the world production has taken place during the last 50 years. It is likely, however, that cadmium constituted a potentially harmful element to exposed humans, occurring in nature together with zinc, lead and copper. The major human health hazard of Cd is a decline in the renal function even at moderate intake (Steinnes, 2009)

Chromium is well known as a toxic metal in the CrVI form, provoking severe metabolic disorders, membrane damage, cancer and contact dermatitis (Bini et al., 2000).

Mercury is a heavy metal ubiquitous in the environment, whose global emissions have been estimated around 4.000t/year, constituting a health risk for the human population; indeed, excessive Hg exposure resulting from burning the Au/Hg amalgam, vapours inhalation, contaminated fish consumption lead to adverse health effects (Thornton, 1996; Steinnes, 2009), causing episodes of poisoning (neurotoxicity, mercurial tremor, psychomotor retardation).

Arsenic is well known as a highly poisonous metalloid, particularly in the water-dissolved form; As concentrations up to 600 µg/L in artesian well waters in Taiwan, India and Bangladesh, for example, have been related to increased risk of skin and internal cancers affecting more than 40 millions people (Steinnes, 2009); a reduction to 10 ppb of As concentration

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Claudio Bini

as a tolerable level in drinkable waters, therefore, has been recently proposed by the World Health Organization (2004).

1.3. CHANCE

As already mentioned, active and abandoned mining activities are widely diffused at worldwide level. Many countries have abandoned mine workings since the ‘70s of last century, and from that time are developing actions to minimize the environmental impact during and after exploitation, and projects aimed at restoration of contaminated/degraded areas. Examples of these research projects are reported in current literature (see f.i. Berger et al., 2000; Mendez and Maier, 2008, and references therein).

Once mines have been closed, and waste abandoned on the land or discharged into surface waters, (since mining industry was completely unregulated until half 19th century, as reported by Davies, 1987), decision makers, after claiming the environmental damage had been done, have discovered that these sites could constitute a challenge, or rather a chance, to rehabilitate the contaminated land. Archaeologists and geologists joined their effort to discover ancient settlements nearby former mining sites, and to understand the organization of ancient societies, their evolution with time, and the metallurgical works that characterized the economy of the interested areas (Francovich, 1985; Costagliola et al., 2008; Dill, 2009). To date, under the stimulus of modern historigraphy that pays particular attention to mining and metallurgical concerns, many mine-archaeological parks have been established, particularly in Europe (France, England, Austria, Germany, Poland). In Italy, where many mining and metallurgical monuments of pre-industrial times are located, studies about this subject flourished since the end of 19th century (see Cipriani and Tanelli, 1983, and references therein). Since that time, many initiatives succeeded, giving a profound insight into historical, archaeological, socio-economical and industrial (metallurgical) aspects of former mining sites (e.g. D’Achiardi, 1927; Francovich, 1985; Tanelli, 1989; Costagliola et al, 2008, and references therein), aimed at the valorisation of the land with the opening of several mine-archaeological parks, recreational itineraries and museums in Tuscany, Sardinia, Veneto (Figure 3).

10

Introduction

Figure 3. Archaeological Mine Park at Rio Elba (Elba Island). Tourists looking for minerals. (Photo Bini).

Archaeological investigations carried out in these areas have lead to discoveries of human activities during three millennia (Casini, 1993). The most significant discoveries are related to the extraction, processing, and commerce of metals. For example, excavation of the ruins of the village of Rocca S. Silvestro (a Middle Age village in Central Italy) suggests that it was inhabited by at least 300 persons, devoted to the processing of lead and copper (Francovich,1985).

Archaeological studies also indicate four major periods of settlement and human activity in the territory (Heimann et al., 1998). First, mining debris and stony artifacts (scrapes, tips) of pre-historical and proto-historical periods (Middle Palaeolithic-Neolithic) have been found close to shelters. Second, excavation of settlements of the Etruscan and Roman period revealed intensive metal mining and limestone quarrying activity. Third, in the Medieval period lead and copper processing proved an important activity at different sites in various countries (Costagliola et al., 2008; Dill, 2009; Forel et al., 2010). Finally, in more recent times (16 th-19th

century), ore exploitation has been carried out by both local population and

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Claudio Bini

foreign people, such as the Germans, as demonstrated by local nomenclature (Francovich, 1985).

Another chance offered by formerly mining sites is the fact that such sites host wild vegetation genetically tolerant to high metal concentrations. According to Baker (1981), plants may be classified into three groups on the basis of their ability to accumulate metals in their aerial parts. Excluders are those plants whose metal concentrations remain unaffected by metal concentration in soils up to a critical level, when toxic symptoms appear.

Indicator plants are those whose metal concentrations reflect those of the related soil.

Accumulator plants have the ability to take up and concentrate metals from soils containing both low and high levels of metals. Among the species that may tolerate high metal concentrations in their tissues, plants presenting exceptional accumulating ability are referred to as hyperaccumulators. More than 400 wild plants have been reported as metal hyperaccumulators (Bini et al., 2000). A well known hyperaccumulator species for Ni, for example, is Alyssum bertoloni (Baker and Brooks, 1989), for Zn Viola calaminaria and several Thlaspi species (Baker and Brooks, 1989), for Pb Brassica napus (Mc Grath, 1995); Calendula officinalis has been discovered to accumulate chromium (Bini et al, 2000), and Pteris vittata arsenic (Bettiol et al., 2010).

The metal-enriched areas, therefore, represent an ideal natural laboratory where to study the processes in order to provide descriptive models of the interactions between the toxic elements, the pedosphere, the biosphere and the hydrosphere (Ritchie, 1994). The assessment of soil contamination has been extensively carried out through plant analysis (Ernst, 1996). Wild and cultivated plant species (catchfly, dandelion, plantain, marigold, willow, common reed, fescue, maize) have been used as (passive accumulative) bioindicators for large scale and local soil contamination (Bini, 2009). Based on current knowledge, in the last decades, attention has been deserved to plants as tools to clean up metal-contaminated soils, and restoration plans have been addressed to these sites, with application of low cost and environmental friendly phytoremediation technologies (Bini, 2009; 2010).

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Chapter 2

PROCESSES OCCURRING AT THE MINE SITES

2.1. WEATHERING OF MINE SPOILS

Weathering of metal sulphides in exogene environment, and the consequent release of pollutants, is the geochemical process responsible for the contamination of former mining areas. The process occurs even in absence of exploitation, when ore deposits are exposed to atmospheric agents, but is particularly environmentally relevant with extensive exploitation, or when mining operations ceased, and uncontrolled mine waste is abandoned on the land.

Although sulphide alteration constitutes a common process in contamination of wide mine areas, various sources may contribute to environmental pollution: acid mine drainage (AMD), flotation tailings, mine dumps, crushing, grinding and milling plants. Wind, gravity, runoff, surface and ground-waters are the agents that contribute PHEs to the environment as soluble, suspended or transported material. The spectrum of mobilized PHEs varies depending on the ore deposits composition: Cd, Pb, Zn are the most common, while As, Cu, Co, Hg, Ni, Sb, Se, Te are less frequent.

Among the mixed sulphide deposits, the most hazardous to the environment are those bearing Fe-, Zn-, Pb-minerals in both the exploited and the raw material (gangue). The alteration of these minerals proceeds

via an oxidation reaction that involves the sulphide or disulphide (transformed in sulphate), together with oxidation of iron to Fe3+, and subsequent hydrolysis to Fe(OH)3. The role of iron in oxidation and hydrolysis is particularly important, given the abundance of Fe-minerals in ore deposits.

Iron sulphide oxidation reactions (mainly pyrite and pyrrhotite), tend to create an acidic environment, releasing protons, as observed in the oxidation reaction of pyrite:

FeS2 + 3,75 O2 +3,5 H2O Fe(OH)3(s) + 2 SO42- + 4H+

Pyrrhotite is a not stoichiometric iron sulphide with different polytypes (Fe(1-x)S, where x = 0 - 0,125). The reaction kinetics depends on pH, temperature and surface area, besides the polytype, being the monocline pyrrhotite more reactive than the hexagonal one (Salomons, 1995).

The oxidation reaction of pyrrhotite can be syntetized as follows:

Fe(1-x)S(S) + (2-x/2) O2 + x H2O (1-x) Fe2+ + SO42- + 2x H+

As it is evident, the higher the S/Fe ratio, the higher is the proton release (i.e. acidification).

Iron sulphides oxidation, as observed by Fanfani (1997), creates an acidic environment that enhances metal mobilisation from mixed sulphides, according to the following reaction:

MeS + Fe2(SO4)3 + 1,5 O2 +H2O MeSO4 +2FeSO4 + H2SO4 ( with Me = Cu, Pb, Zn…)

To the acidification process contribute all the mixed sulphides (Fe, Cu, Zn, etc.): for example, in the case of chalcopyrite the reaction is:

4CuFeS2 + 17O2 + 6H2O 4FeOOH + 8SO42-+ 4Cu2+ + 8H+

Processes Occurring at the Mine Sites

Oxidation of not-iron bearing sulphides, as sphalerite or galena, runs in such a way that base metal sulphates form, but no acidification occurs:

MeS + 2O2 Me2++ SO4 2- MeSO4 (with Me: Cu,Pb,Zn…)

Instead, if iron sulphides are present in the environment, as it happens frequently, hydrogen ions are produced, according to the reaction:

MeS+8Fe3++4H2OMe2++8Fe2++SO42-+8H+.

2.2. ACID MINE DRAINAGE (AMD)

As already mentioned, iron sulphide oxidation produces acidic drainage water (AMD). Prediction of AMD is the key factor in predicting the release of dissolved metals from active and past mining operations ( Salomons, 1995; Moncur et al., 2009). The prerequisite for AMD is the generation of protons at a faster rate than it can be neutralised by any alkaline materials in the waste (e.g. carbonate in the gangue), the access of oxygen and water, and a rate of precipitation higher than evaporation.

The most common mineral causing AMD is pyrite, but other metal sulphides may also contribute. The oxidation of pyrite, preceding iron hydrolysis, occurs in three steps. The first one occurs at pH above 4.3, with high sulphate and low iron concentrations, with little or no acidity, and slowly.

FeS2+7/2O2+H2O=Fe2++2SO42-+2H+

The reaction may proceed both abiotically and by direct bacterial oxidation (Lindsay et al., 2009).

The second step occurs with a pH range between 2.5 and 4.15; there are high acidity and total iron increases. The Fe3+ /Fe2+ ratio is still low:

Fe2+ +1/4 O2 +H+=Fe3+ + 1/2H2O

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Claudio Bini

This stage proceeds predominantly by direct bacterial oxidation determined by the activity of microorganisms of the genus Acidithiobacillus.

The third step occurs at pH values below 2.5, high sulphate and iron levels. The ratio of Fe3+/Fe2+ is high. The reaction is totally determined by bacterial oxidation, that enhances solubilisation of metal sulphides, catalyzed by chemolithoautotrophic acidophile microorganisms (e.g. Thiobacillus Ferrooxidans) (Trois et al, 2007):

FeS2 + 14 Fe3+ + 8H2O=15 Fe2+ + 2 SO42- + 16H+

These three stages are the primary factors, directly involved in the acid production process (Ferguson and Erickson. 1988). The intensity of acid generation by these primary factors is determined (Salomons, 1995; Fanfani, 1997) by environmental (e.g. grain size, pH, temperature, oxygen concentration, metal activity) and biological parameters (population density of the bacteria, rate of bacterial growth, supply of nutrients).

Secondary factors control the consumption or alteration of the products from the acid generation reactions. Neutralisation of AMD can occur when an effective buffer system with relatively high pH is established, thus impeding Fe(III) mobilization (Fanfani, 1997). This occurs when carbonate minerals (calcite, dolomite or ankerite) are present. At pH <7.2, the carbonate-bicarbonate equilibrium turns towards bicarbonate:

CaCO3 + H+ HCO3- + Ca2+

.

There occur four moles of carbonate to neutralize one mole of pyrite; pH is buffered at a range between 6.4 and 5.5 (Ritchie, 1994), a value at which iron in oxidizing environment precipitates as oxyhydroxide or as sulphate (jarosite):

4FeS +8CaCO3 + 6O2 + 4H2O 4FeOOH + 4SO42- + 8Ca2+ + 8CO2

Neutralisation by carbonates is a relatively fast process and provides short-term buffering capacity.

Other buffering systems, as Fe and Al hydroxides, or silicate minerals, operate at much lower pH, and do not prove effective in controlling the

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metal release from sulphides, providing long-term buffering capacity. However, AMD neutralization could be not sufficient to eliminate contamination, since sulphides oxidation, although slowed down, is still active, and the release of pollutants simply occurs later than at lower pH. Submersion of waste in artificial impoundments with anoxic conditions could be an effective technique to prevent AMD production and pollutant release in the environment.

2.3. FLOTATION TAILINGSX DELETE

As already mentioned, AMD contains elevated levels of metals. One way to attenuate, although not eliminate, environmental pollution by mine waste, is AMD neutralization. This may be attained through different methods and techniques. Yet, there are several physical, chemical and biological processes operating in the natural environment, that can contribute to contaminant attenuation. Physical processes include: physical mixing of waste particles with uncontaminated eroded soils and sediments particles; proportional dilution and dispersion of pollutants during high discharge and surface run-off, and metal confinement and sedimentation into confined basins (Figure 4). Chemical processes include solution (metal-soluble fraction), complexation (organic matter-bound fraction), precipitation (oxide-bound fraction), and adsorption by suspended particles (exchangeable fraction). These consist mainly of clay minerals, iron oxyhydroxides and organic matter (Salomons, 1995).

Surface erosion by water or wind, or direct discharge of waste materials in rivers, may result in the introduction of metals in particulate form into aquatic ecosystems, and the heavy metals can be transported considerable distances downstream, causing extended contamination. The leaching time of sulphides from oxygenated spoils is estimated to be about 11 years on average (Helios-Rybicka, 1996). The water discharge supplies 100 tonnes of total dissolved soils per day, with base metals (mostly Pb and Zn) up to 2gm-3 n(Helios-Rybicka, 1996). However, some spoil dumps can be persistent sources of contamination with products of sulphide oxidation, which may affect the environment for decades.

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Claudio Bini

Figure 4. A typical flotation basin in the Sulcis mine district, Sardinia, Italy. (Photo Bini).

Initially, in former mine works, AMD was convoyed to nearby streams, increasing metal concentrations in water and overbank sediments for more kilometres downstream. Successively, although too late to avoid environmental damages (in particular to the aquatic ecosystem) (Davies, 1987), owing to the major sensitivity of population, apposite flotation basins were built, with the goal to limit water and surface contamination. Flotation impoundments have been long utilized to reduce the environmental impact of fine particles produced by metal processing, in such a way that pollutant attenuation may occur.

When the AMDs reach the impoundment, a wider dispersion of the metals both in solution and (after adsorption) in particulate form is possible. Benvenuti et al. (1999) found that there is a gradation in grain size from sand and silt close to the mouth of the drains, to mud and clay at the opposite site; during wet season, the whole impoundment may be flooded, and tailings alteration is less pronounced, presumably because oxidation of the waste is impeded by water saturation in reducing conditions (Neel et al., 2003), when Eh drops to <250mV.

A similar gradient in metal concentration was observed, due to dilution connected to adsorption/precipitation processes occurring in the flotation basins (Benvenuti et al., 1999). Near the dumps, the water is acid (average pH 3.0) and tailings have an high metal content (average Pb 369 ppm;

18


average Zn 176 ppm); the pH increases dramatically (up to 8.0) with increasing distance from the dumps, and the metal content decreases so far. Berger et al. (2000) point out that natural attenuation in drainage from a historic mining district may be related to two distinct pathways: metals (Al, Cu, Fe, Pb) precipitate directly from carbonate-rich solution, whereas Zn, Mg, Mn and SO4 concentrations decrease primarily through mixing (i.e. dilution) with tributary streams.

Sulphide weathering (oxidation, adsorption or coprecipitation by iron hydroxides) was identified in tailing ponds in the unsaturated proximal areas beside the earthen dams (Heikkinen and Raisanen, 2009) at an active mine site in Finland, where the raised water table contributed to desorption and remobilization of metals, probably through dissolution of iron precipitates.

Sequential extractions applied to mine tailings (Fanfani et al., 1997; Conesa et al, 2008; Perez-Lopez et al., 2009) showed that a relevant part of the total amount of metals convoyed in flotation basins (around 90-100% of total S, Zn, Co and Ni, 60-70% of Mn and Cd, 30-40% of Fe and Cu, and 5% of As and Pb) was estimated to be in the bio-available fraction, i.e. potentially harmful to the aquatic ecosystems.

Leaching experiments carried out by Da Pelo et al. (2009) in Sardinia mining sites, and by Palumbo-Roe et al. (2009) on mine tailings in Wales, show that where surface waters interact with mineral assemblages of the alteration zone, this corresponds to a marked increase in pH concomitant with a decrease in dissolved metals. A comparative slow reaction rate results in the release of a harmful amount of contaminants (Musu et al., 2007). Solute transport in the tailings is governed by unsaturated flow and is controlled by the seasonal precipitation–evapotranspiration cycle. It is envisaged that the seasonal movement of the saturated/unsaturated surface in the tailings in response to seasonal capillary pressure changes is responsible for causing the solute transport. The results of the percolation tests are consistent with control of metal concentrations by mechanisms of dissolution/precipitation/sorption, whereas there is no evidence of sulphide oxidation during the leaching.

The percolation test best describes the seasonal flushing of the secondary minerals, products of metal sulphide oxidation, from the surface layers of the tailings, whereas it does not address the sensitivity to redox

19

Claudio Bini

changes of the waste. This aspect becomes significant during periods of exposure of the tailings to alternating wet and dry periods.

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

20


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).

21

Claudio 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).

22


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).

23

Claudio 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;

24


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.

2.6. BIOLOGICAL IMPLICATIONS

There is a general, although not simple, relationship between the heavy metal content of soils and plants growing on those soils (Davies, 1987; Baker and Brooks, 1989; Adriano, 2001). Uptake through the roots is influenced by soil parameters such as acidity or redox potential, and different plant species absorb metals to different extents. But in general, the higher the heavy metal concentration of the soil, the higher will be the concentration in plants (Baker, 1981).

An absorption sequence Zn>Co>Cu>Ni>Fe>Cr, consistent with leaching tests, was found by Dinelli and Lombini (1996) in wild plants growing on mine soils, suggesting plant uptake to be controlled by soil solution composition. Fontana et al. (2010) report that in wild plants of mine soils the less mobile among the trace elements considered is Pb (average TF = 0.37), which tends to remain blocked in the roots, because it is not essential for plant nutrition, thereby suggesting some exclusion strategy by plants. Chromium, Cu, Zn, Fe, are present in similar concentrations in leaves and roots (TF ≈ 1), while Mn appears to be the most translocated among the elements considered (average TF = 2.33). metal translocation is probably influenced by the bioavailability of the metal and by the species of plant considered, which are two determining key factors in the evaluation of the absorption of pollutants by the plant compartment (Kabata-Pendias, 2004). Similar results were obtained by Bini et al (2000) for chromium in wild plants growing on Cr-contaminated soils.

Among heavy metals, cadmium and lead are toxic to animals and people, and the accumulation of these metals in foodstuffs raises the question whether human health might be impaired by ingesting small amounts of toxic metals. Thornton (1996) reported lead concentrations in garden vegetables in mining areas of Derbyshire to be 2-4 times higher than those of urban soils; accidently ingested soil resulted to be the major

25

Claudio Bini

source of Pb intake. In an early paper, Davies and Roberts (1975) found that radish plants (Raphanus sativus L.) from some gardens close to mine sites in Wales, contained more than the British legal limit for lead in food (2mg/kg fresh weight), and cadmium concentrations were high enough to cause concern. Similarly, mining at a site in SW England resulted in extremely high Cd (and Zn) concentration in soils and leafy vegetables, including cabbage, spinach and lettuce (Thornton, 1996).

Exposure to heavy metals may affect severely human population by metabolic and neurological disorders, psychomotor retardation, intoxication, respiratory diseases, liver and kidney damage, skin and internal cancer (Jarup, 2003). Acute toxicity by lead inhalation or ingestion, both direct and indirect, via the food chain (Abrahams, 2002), brings out malfunctioning to the reproductive system, kidney insufficiency, damage to neurological system and brain. Food ingestion has been found by Zheng et al. (2007) to be the most common way (>90%) of human exposure to metal contamination, in comparison to other exposure ways such as inhalation and dermal contact.

26

Chapter 3

CONSEQUENCES OF MINING OPERATIONS ON ENVIRONMENTAL TRANSFORMATIONS

AND MINE SOIL EVOLUTION

Numerous processes, both physical and chemical, contribute to environmental transformations consequent to former mining activities and subsequent restoration of the exploited areas.

In the initial stage of rock alteration, physical processes prevail. These are particularly effective in areas with steep morphology, where most mining districts are located. Loose and coarse grained material forms as a consequence of rock fragmentation. Rock fragments migration on instable slopes, erosion of fine particles by wind and runoff, all these processes contribute to land modelling. Meanwhile, chemical processes begin to act, further contributing to environmental transformations by oxidation (Eh>250mV), acidification (pH<7), hydrolysis, metal leaching, precipitation of oxyhydroxides and sulphates, argillogenesis. All these processes may be conditioned by water availability and temperature (i.e. the climate conditions), that enhance mine waste reactivity.

As a consequence of the above processes, a set of physical and chemical features characterize the soils developed from mine spoils. Once the parent material is finely subdivided and weathered, the formation of a biologically active substrate may occur, thereby permitting pioneer

vegetation (lichens, mosses) development (Burykin, 1985). Litter accumulation (OL horizon) is the process that characterizes the early stage of soil formation. Subsequently, organic matter decomposition (OF horizon), humus formation (OH horizon), mineralisation (A horizon) constitute a first pedogenetic phase. According to Jabiol et al. (2007), this phase may bring to the differentiation of several types of humus as a function of litter composition, microflora and microfauna activity, pH and climate conditions.

A second pedogenetic phase is determined by in situ mineral transformations (e.g. acid hydrolysis), oxyhydroxides and clay formation (stage of cambic horizon formation). In this phase, colour varies from very dark brown (10YR 3/3) to dark brown (7,5YR 3/3), reddish brown (2,5YR 3/2), dark yellowish brown (10YR 4/5), or blackish (5YR 2,5/5), in relation to the nature of the bedrock, and /or to the amount of mine waste.

A third pedogenetic phase is consistent with solute leaching and particles migration towards bottom (stage of argillic horizon formation); precipitation of new minerals (e.g. carbonate, sulphate) is likely to occur. However, this third phase is difficult to assess in mine waste materials, since the time elapsed from mining operations generally is not sufficient for Bt formation, if we consider that the landscape morphology is generally undulated, with slopes ranging from 15% and 45%, and therefore erosion is a prominent process. Yet, soils developed from waste dumps are generally shallow (20-100 cm), skeletal, coarse-textured (sandy loam to loamy sand), little developed, with limited horizonation.

Table 2. Selected properties of the Spolic Xerorthent illustrated in Figure 8.

Soil Horizon

Depth Particle size (USDA)

%

pH Total Carbonates

Organic Carbon

Organic Matter

CEC

cm silt clay sand g/kg g/kg g/kg cmol(+)/kg

A1 0-10 25 14.8 60.2 7.8 0 16 28 20A2 10-30 26 27.4 46.6 7.6 0 14 24 22.52C 30-70 32.7 37.5 29.8 7.5 0 5 9 14.5

Consequences of Mining Operations ...

3.1. MINE SOILS

Mining activity, although discontinuous in space and time, determined the dispersal pattern of mine spoils. These anthropogenic deposits were irregularly spread over the land surface, giving origin to a new (soil) parent material, completely different from the one the original soils had developed from. Immature Entisols (Lithic Spolic Xerorthents and Spolic Xerorthents as proposed by USDA Soil Taxonomy,1999, and ICOMANTH; Figure 8) are formed in mine spoil <100 y of age. These soils are characterized by a thin solum (<30cm), little organic matter accumulation (mean 14g/kg organic carbon, range 1-33 g/kg), dark brown (10YR3/3) to reddish (5YR4/6) colour, coarse texture (sandy loam to loamy), and subalkaline pH (mean 7.4, range 6.9-7.8). A consequence of the high content of toxic heavy metals, in combination with reduced soil thickness, leads to discontinuous vegetation coverage that is composed mainly of crust lichens, mosses and fescue. Detailed descriptions of soil properties of a selected profile on mine waste are given in Table 2.

Figure 8. Soil profile of a Spolic Xerorthent developed from mine spoils at Temperino, Tuscany. (Photo Bini).

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Claudio Bini

The soils developed from old mine dumps, or in the proximal parts (<0.5 km) of the dumps, are characterized by a solum >50 cm thick, sandy loam to loam texture, blocky structure, slightly acidic pH (mean value 6.3, range 4.9-7.7), humus accumulation (up to 14% organic matter in the A horizon), moderate to low cation exchange capacity (mean 20 cmol(+)/kg), with significant desaturation (base saturation <60%). Generally, they have distinct A-B-C horizonation and a well formed cambic horizon. Therefore, they are Inceptisols (Spolic Haploxerepts or Spolic Dystroxerepts, see Table 3 and Figure 9). Frequently, a discontinuity occurs between the upper part and the lower part of the profile, which developed from the underlying bedrock. Data (not reported) indicate relevant differences and a remarkable polycyclic evolution, owing to the superposition of mine spoil over the normal soil. Colour, texture, reaction, and cation exchange capacity are the most prominent features that present major differences. Soil horizons show dark brown (7.5YR3/2) to dark reddish brown (5YR 3/3) colour, well individualized structure, from crumby to fine blocky peds. Texture is coarse (sandy loam to loam) in surface horizons sampled on mine spoils and loamy to clayey underneath. Soil reaction is slightly acidic (pH 6.3) at the surface, subalkaline (pH 7.4) and base-saturated at the bottom. Cation exchange capacity increases with depth, from 15 to 25 cmol(+)/kg.

Table 3. Selected properties of the Spolic Dystroxerept illustrated in Figure 9

Soil

Hor

izon

Dep

th

Parti

cle

size

(U

SDA

)% pH To

tal

Car

bona

tes

Org

anic

C

arbo

n

Org

anic

M

atte

r

CEC

Exch

. Aci

d.

Bas

e Sa

tura

tion

cm silt clay sand g/kg g/kg g/kg cmol(+)/kg

cmol(+)/kg

%

A1 0-47 34.7 15.3 50 6.4 0 27 46 25.5 15.6 62A2 47-70 33.3 15.1 51.6 6.3 0 21 36 15.4 17.4 47Bw 70-90 30.8 11.1 58.1 6.2 0 8 13 13.0 9.3 58

30

Consequences of Mining Operations ...

Figure 9. Soil profile of a Spolic Dystroxerept developed from mine spoils at Temperino, Tuscany. (Photo Bini).

Figure 10. Soil profile of a Spolic Haploxeralf developed from mine spoils at Temperino, Tuscany. (Photo Bini).

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Claudio Bini

Table 4. Selected properties of the Spolic Haploxeralf illustrated in Figure 10

hori-

zon

dept

hParticle size USDA

pH Car

bona

te

O. C

.

Org

anic

Mat

ter

CEC

cm silt clay sand g/kg g/kg g/kg cmol(+)/Kg

A1 0-3 23,9 9,5 66,6 7,6 0 23 40 29,5A2 3-15 29,8 13,7 56,5 7,7 0 19 33 37E 15-40 35,7 9,9 54,4 7,9 0 6 11 33Bt1 40-110 40,7 40,3 19 8,0 4,1 7 12 19Bt2 110-120 39,9 30 30,7 8,0 58 11 19 23

Shrubby vegetation with shallow trees (holm-oak, strawberry tree, heath, etc.) is the typical vegetation cover at these sites, where rock-rose is the dominant plant.

Soils described and sampled at major distance (>0.5km) from the mine dumps present little evidence of mine spoil in the profile. Sulphidic minerals are found especially at the surface, as revealed by mineralogical and chemical composition (Baldini et al., 2001). An abrupt textural change (Table 4) indicates a marked discontinuity between the upper and lower part of the soil profile. The upper part (A and E horizons) has dark brown (10YR2/2) to yellowish brown (7.5YR3/4) colours, loam to sandy loam texture, crumby structure, high organic carbon content (mean 21 g/kg), and subalkaline pH. The lower part (Bt horizon) presents reddish colours (5YR3/4 to 2.5YR3/4), a strong clay content increase (clay loam to clayey texture), organic carbon decreases (as expected), pH is subalkaline with traces of carbonate. These features are consistent with soil development from limestone in the Mediterranean environment. Therefore, they are classified as Alfisols in the USDA Soil Taxonomy (1999). Since there is evidence for mine waste in the profile, these soils should be classified as Spolic Rhodoxeralfs or Spolic Haploxeralfs (Figure 10). However, considering the net discontinuity already mentioned, these soils could be classified as Spolic Xerorthent over Typic Rhodoxeralf (or Haploxeralf).

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Chapter 4

CASE STUDIES IN ITALY

Besides early investigations of the great Italian geologist Bernardino Lotti (1847-1933), previous studies carried out by several working groups on mine sites in Italy ( Zucchetti, 1958; Burtet Fabris and Omenetto, 1971; Corsini et al., 1975; Zuffardi, 1977, 1990; Gianelli and Puxeddu, 1978; Lattanzi and Tanelli, 1981; Cipriani and Tanelli, 1983; Deschamps et al., 1983) pointed primarily at understanding the complex genesis of ore deposits and the possibility of mineral exploitation (Figure 11). More recently, after the closure of mines, attention was focused on the environmental impact of mining operations (Leita and De Nobili, 1988; Benvenuti et al., 1999; 2000; Caboi et al., 1999, Mascaro et al., 2001b; Bini and Gaballo, 2006; Cidu et al., 2009; Fontana et al., 2010), and possible land restoration (Dinelli e Lombini, 1996; Zerbi and Marchiol, 2004; Marchiol et al., 2010; Bini et al, 2010; 2011). In fact, once ore deposits were exploited, environmental problems connected to the discharge and spreading of mine waste on conterminous land and streams became a concern, constituting elsewhere a waste area on the modern landscape. The tailings surface remained uncovered for a long time, until weathering and pedogenic processes enhanced soil formation and revegetation in the mine spoil, producing significant environmental impacts in the whole mining area (Benvenuti et al., 1999).

A soil survey of the abandoned mine areas in Italy has been on-going since the 1990s at various Universities (Cagliari, Florence, Milan, Siena, Udine, Venice) and Research Centres, in the frame of a national research project. The survey was preceded by mapping the distribution of mine

spoils discharged at the surface. The rationale for soil sampling was focused on soils developed from mine spoils of different age and in the conterminous areas with soil unaffected by spoil. Several soil pits were opened and profiles (totally more than 200 pit soils) described and sampled in the following type of sites: spoil areas, no spoil proximal areas (spoil <0.5 km), and distal areas (spoil> 0.5 km). At some sites, roastings, flotation tailings and overbank sediments were sampled. The results are summarised in the case studies hereafter.

Figure 11. Regions of Italy. Numbers are abandoned mine sites cited in the study: 1 = Elba Island; 2 = Metalliferous Hills; 3 = Temperino mine; 4 = Bottino, Apuane Alps; 5 = Sardinia, Sulcis-Iglesiente district; 6 = Imperina Creek Valley, Dolomites.

6

5

1 2

4 3

Case Studies in Italy

4.1 TUSCANY

There is a long history of mining activity for mixed sulphides (mainly Cu, Fe, Pb, Zn) at several sites in Tuscany. Mining dates back at least to Etruscan times (7th cent. B.C.), flourished under the Romans (1th cent. B.C.), during the Middle Age-Renaissance (10th-16th cent. A.D.), and in the 19th - 20th century (Tanelli, 1985).

The southern Tuscany metallogenic province (including the Elba island) is of primary importance due to the occurrence of several ore deposits associated with volcano-sedimentary, magmatic, metamorphic and geothermal environments (Lattanzi et al, 1994; Costagliola et al., 2008). The mining district is characterized by deposits of pyrite and mixed sulphides (Fe, Cu-Pb-Zn-Ag, As, Sb, Hg, Sn and Au).

Among the many types of ore deposits occurring in southern Tuscany, the Fe oxide deposits of Elba island and the pyrite and other base-metal sulphide deposits of the Metalliferous Hills district have been extensively exploited since the 1st millennium BC under the Etruscans, although metal mining and smelting dates back to the late Bronze age (Cipriani and Tanelli, 1983; Corretti and Benvenuti, 2001).

4.1.1. Elba Island District

Elba Island was one of the most important Italian mining sites, dating back to Roman age, as it is demonstrated by metallurgic findings of Roman period (Costagliola et al., 2008). Iron exploitation in the island ceased in the 1980s (Servida et al., 2009), and a part of the ancient mining area is presently used as an open site, where tourists may search and collect minerals (see Figure 3). More than one hundred mineralogical species have been recorded in the various ore deposits of the island. The main ore bodies of Elba island are located in a narrow belt along the eastern coast of the island, where they occur in variable settings (lodes, veins, pods), differently related to the host rocks (Costagliola, 2008). The primary mineralogy of the ore deposits is composed mainly of Fe-oxide (hematite, limonite and magnetite) in the northern part, whereas iron and base-metal sulphides (pyrite, chalcopyrite ±As, Bi, Pb, Sb, Zn) are more common in the southern part. More than 4.46x106 m3 of material was removed (Servida

35

Claudio Bini

et al., 2009), and most minerals exported and smelted at smelting centres in Southern Tuscany (Costagliola et al., 2008), particularly during Etruscan and Roman periods.

The accumulation and the exposure to the atmospheric agents of the sulphide-bearing earth materials without adequate management initiated the AMD processes The consequent environmental hazard depends on the mobility of metals, the climatic conditions, the porosity of earth material, etc. (Servida et al., 2009).

The average content of selected heavy metals, pH and texture of earthy material in the vicinity of mine sites at Elba island is reported in Table 5. Bulk elemental composition of mine waste is similar to silicatic bedrock and related soils. Metal amounts are higher in the close vicinity of mines, and decrease with distance from the mine waste. Correspondently, the pH shows an opposite trend, increasing with distance, while texture decreases. It is likely that a geochemical halo forms around the metal hotspot, and a dilution occurs with distance from the mine waste. Iron is the most abundant metal, as expected from the geology of ore deposits, and decreases sharply in distal soils; basic metals, instead, are rather persistent in the examined soils.

Table 5. Average trace elements concentrations, mean values of pH and texture at different sites in the Elba mine area

Fe Mn Cu Pb Zn pH texture

Mine waste

22 0.2 730 336 919 n.d. gravel

Mine soil (1995)

18 0.09 230 146 634 5.3 Gravel, sand

Proximal soil (1995)

12 0.07 120 164 451 6.3 Sandy loam

Distal soil (1995)

8 0.07 62 131 401 6.6 Loamy sand

Source: Corsini et al., 1980; Bini et al., 1995; Servida et al., 2009).Fe, Mn are expressed as %; Cu, Pb, Zn as mgkg-1.

4.1.2. The Metalliferous Hills District (Massa Marittima)

36


A number of polymetallic Cu-Pb-Zn-Ag vein deposits, controlled by late Apenninic horst-and -graben structures, have been mined for several millennia in the area around the medieval town of Massa Marittima (Benvenuti et al., 1999). Yet, the Massa Marittima area hosts mineral deposits of variable extent and importance, located at different sites (Costagliola et al., 2008). In the past century exploitation focused mainly on the Cu-Pb-Zn-Ag deposits of Fenice Capanne, and the pyrite deposits of Niccioletta, Boccheggiano and Gavorrano. Minor mineral occurrence was located at Monte Arsenti, where Cu-Zn-Pb-Ag sulphides and sulphosalts were exploited discontinuously in the past 3000 years (Cipriani and Tanelli, 1983).

Mine ores, including those brought out from the Elba island, were smelted at different metallurgical centres in the close vicinity of Massa Marittima, where stream water, wood, and refractory material (clay and sandstone) were easily available (Costagliola et al., 2008).

The techniques for separating ore from rock remained broadly the same throughout a long mining period. At first, smelting took place near the mines but diminishing supplies of wood for burning and roasting led to the development of specialised smelting centres (Davies, 1987). Massa Marittima was one of the most important mining and metallurgical centre where Pb, Cu, (Ag) production was carried out mainly during the Middle Age, in the XII-XIV century (Mascaro et al., 2001), and under the Medicean dynasty (Cipriani and Tanelli, 1983), leaving about 500.000metric tonnes of Fe slags, dismantled and re-used in the last century for reclamation of surrounding land (Costagliola et al., 2008).

4.1.3. Fenice Capanne

The Fenice Capanne sulphide deposit is mainly constituted of two polymetallic vein bodies, linked to the principal tectonic dislocation in the area (Benvenuti et al., 1999). The primary mineral association is characterised by the presence of predominant chalcopyrite (Cu 2.5-11% d.w.) with Zn and Pb sulphides (Zn 4.5-16%; Pb 1-3.5%). The ore body originated from hydrothermal processes associated with the late orogenetic magmatism, which were responsible also for the alteration and replacement of host rocks, with the formation of skarn silicates (mainly pyroxenes,

37

Claudio Bini

epidotes, ilvaite), pyrite, chalcedony, kaolinite and alunite (Mascaro et al., 2001).

At Fenice Capanne, exploitation and processing of the polymetallic (Zn, Cu, Pb, Fe, Ag) sulphide deposits, initiated during Etruscan times (6 th-7th century B.C.), closed in the 1980’s, being the last operative mine in Southern Tuscany. The main exploitation occurred in the medieval age and during the 19-20th centuries. Total production is estimated in the order of some thousand tonnes of Cu, Pb, Zn and some tonnes of Ag (Mascaro et al., 2001a).

The morphology of the area is gently undulating, with maximum elevation around 450 m a.s.l.; vegetation is represented by Mediterranean maquis. Climate is mesothermic subhumid with marked summer deficit (maT = 15°C, maP = 850 mma-1 ). The river Bruna (20km length) is the main waterway draining the whole area (approx 35 km2 ).

The mine waste produced by metal exploitation includes huge dumps of roasting products and excavation waste, mainly dating back to the 19 th

century, and flotation tailings produced during the period 1950-1984. The oldest waste were dispersed on the surrounding land, while roasting dumps occupy about 250ha, partially forested.

The roasting piles are produced by the roasting of low grade ore (<4% d.w. Cu). Most excavation and roasting dumps are 15-20 m high and partially reforested. The flotation tailings are mainly discharged into four artificial impoundments filled in the period 1957-1984. Their total capacity is about 850,000m3. The oldest flotation tailings lie over an unconfined area, spatially associated with piles of roastings.

The particle size of waste material is variable: roasting products are composed of coarse sand and gravels; the pH is acidic, in the range 3.3-3.8. Unsaturated water conditions, and the coarse texture of roastings, enhanced sulphide oxidation; therefore, secondary minerals (jarosite, barite, gypsum) formation occurred.

The flotation tailings, on the contrary, are fine-grained (silty-clayey), with slightly acidic to neutral pH (range 6.2- 7.8). Secondary mineralogical phases are Fe-oxyhydroxides, jarosite, gypsum, illite, kaolinite. Water draining flotation tailings 1km downward present acidic pH and high metal and sulphate contents.

The mineralogy of mine waste has been recently studied in detail by Mascaro et al. (2001a). Quartz and hematite, partly derived from roasting

38


processes, are the most abundant primary mineralogical phases (see Jambor, 1994, for definition). Minor contents of feldspars, epidotes and phyllosilicates (muscovite and clay minerals) are present, together with pyrite, traces of chalcopyrite, barite and galena. The secondary phases are mainly Fe oxyhydroxides, jarosite and gypsum. The flotation tailings consist mainly of quartz and feldspar, with minor hedenbergite, phyllosilicates, hematite and pyrite; sphalerite and chalcopyrite are present only in traces, whereas carbonates (dolomite, siderite, calcite) are always nearly absent. The main secondary phases are still Fe oxyhydroxides, jarosite, gypsum and clay minerals. Sulphides, siderite and dolomite are replaced, partially or completely, by Fe oxyhydroxides. Microscopic observations (Mascaro et al., 2001a) show that feldspars are mainly orthoclase and rare adularia, with minor contents of plagioclase; commonly, silica is represented by chalcedony.

All types of mine waste materials show in the fine grained fraction high contents of illite and kaolinite (about 70-100% of total clay minerals). Other clay minerals are chlorite, montmorillonite and illite-montmorillonite interstratified. Kaolinite probably occurs both as primary hydrothermal phase and secondary phase: SEM-EDS observations show the formation of secondary kaolinite from the alteration of primary silicates (in particular K-felspar) in more acidified mine tailings.

Most of the samples have high and variable contents of toxic elements, in particular Pb, Zn and Cu (Table 6). The metal content commonly decreases with decreasing age of waste; proximal and distal soils show nearly the same concentrations, suggesting limited dispersion to occur with distance. Zinc constitutes an exception in this frame. The near neutral pH suggests that the processes of sulphide oxidation and of acid solution buffering balance each other. The acidification and sulphide oxidation observed in the roastings is probably caused by the lack of carbonates and by the large grain size and long residence time.

Soils proximal to mine waste and roastings are scarcely deep and present limited development, with simple A-C horizonation. Texture is sandy skeletal, reaction acidic (pH=3.3-3.8); the ratio C/N=14 suggests humification to be effective in surface horizon; bulk density is low (1.2gcm-3).Table 6. Average concentrations of heavy metals, pH and grain size at

Fenice Capanne

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Claudio Bini

Sample (age)

Fe % As Bi Cu Mn Pb Zn pH Texture

Roastings (ca.1890)

17 927 320 3,450 418 825 2,480 3.5 Gravel, pebble

Tailings (ca.1950)

3.6 515 22 853 424 2,050 2,490 4.2 Sand, silt

Tailings1 (ca.1960)

7.7 230 84 3,400 4,320 1,310 5,470 7.1 Sand, silt

Tailings2 (1970)

5.1 116 30 508 3,520 214 1,560 7.4 Silt, clay


5.8 123 28 595 3,700 353 1,720 6.4 Coarse loamy

Distal soil (1999)

6.3 119 27 445 2,900 413 522 6.7 Coarse loamy

Adapted from Mascaro et al., 2001a.

Distal soils too present limited differentiation, but have silty-clayey texture, neutral pH (6.2 – 7.8), C/N ratio = 15, bulk density a little higher (1.4gcm-3). The mineralogy of the fine fraction is composed of illite, kaolinite, jarosite, gypsum, Fe-Mn oxyhydroxides.

4.1.4. Boccheggiano

The Boccheggiano district was one of the most important mine areas in Italy, and has been widely explored since the early ‘900 (D’Achiardi, 1927).

Mining activity in the Boccheggiano area of southern Tuscany has been documented back to the 16th century A.D., but likely dates to at least Etruscan times (Benvenuti et al., 1997; 1999). Since than, a number of base metals (Hg, Sb, Fe oxides, pyrite, chalcopyrite) have been intensely exploited in the district, yielding about 1.5x106 tonnes of ore at 4-8% Cu in the last century. Up until about 1910, the main focus was on base-metal (Cu(Pb-Zn-Ag)) ores, but then, from 1906 to 1994, some tens of millions

40


of tonnes of pyrite were produced from several deposits in the area (Tanelli, 1985; Lattanzi et al., 1994).

The geology of the area is composed of Palaeozoic metamorphic basement in strict contact with the Mesozoic Tuscan nappe and Cretaceous-Eocene flysch; the morphology is undulated, with elevation up to 700m a.s.l. and 35-40% slope. Climate is warm Mediterranean (maT=16°C, maP=1020mm); the land use is partly a mixed forest with prevailing oak, and partly arable land; part of the area was afforested with pine (Pinus pinea L.) in the ‘50s.

The Merse river drains the region with mean flow of 7m3sec-1 and frequent flooding episodes.

Such an extensive and protracted mining activity has left behind many abandoned mines and mine wastes, and huge masses of slags and roastings. The dumps of copper-pyrite excavation waste are 10-15 m high, and extend for about 1.5 km along the Merse river.

Three main types of mine waste have been identified in the study area: waste-rock dumps, a flotation tailings impoundment and roasting-smelting waste (Benvenuti et al., 1997; 1999). Waste rock material of mining activities represents a primary source of pollution for drainage waters, sediments and soils, because of the generally high metal concentrations. Benvenuti et al. (1997, 1999) studied the mineralogical and chemical features of mine tailings sediments, soils and drainage waters, with special focus on the exogenous minerals considered mineral traps for the toxic elements (Cu, Pb, Zn, As, Cd, Bi), and on the dispersion mechanisms and halos of these elements.

Following Jambor suggestions (1994), in the dumps the Authors distinguished three classes of minerals: 1) primary (ore, gangue, and pyrometallurgical phases: sphalerite, galena, pyrite, chalcopyrite, iron oxides, quartz, calcite, micas, chlorite); 2) secondary (minerals formed in situ within the waste disposal area): Fe, Al and Cu sulphates (jarosite, copiapite, alunite, chalcantite), gypsum, Cu carbonates (malachite), Fe oxyhydroxides, and 3) ternary and quaternary minerals, not formed in situ, but developed after sampling and oven-drying at >60°: trace amounts of siderotil, bassanite, metaluminite.

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Table 7. Average and range of trace elements concentrations and mean values of pH and texture at different sites in the Boccheggiano

mine area

Sample (age)

Fe As Bi Cu Pb Zn pH Grain size

Mine waste (ca.1889)

12.8 (1.6-19.3)

188 (13-264)

153 (37-339)

589 (529.642)

590 (277-1110)

292 (234-327)

3.1 Gravel, pebble


13.6 (4.7-24.5)

241 (88-429)

26 (4-80)

754 (36-1790)

3160 (34-28000)

758 (77-1970)

4.2 Gravel, pebble


10.4 (9.4-11.3)

737 (55-1000)

29 (14-40)

196 (29-468)

304 (106-956)

392 (40-1300)

4.3 Pebble, sand

Tailings (1970)

9.4 (5.2-15.2)

233 (74-887)

13 (3-44)

203 (71-377)

2145 (449-4920)

4980 (154-9900)

5.5 Silt, clay


10.1(1.6-15.9)

640 (50-900)

17 (2-29)

81 (35-800)

402 (110-956)

312 (97-446)

4.8 Coarse loamy

Distal soil (1995)

4.2 (1.6-5.8)

55 (40-350)

10 (1-14)

59 (29-206)

81(50-890)

176 (40-331)

6.7 Coarse loamy

Fe is expressed as %; As, Bi, Cu, Pb, Zn as mgkg-1. Adapted after Benvenuti et al., 1999.

Waste samples show extremely variable amounts of metals, and this feature may be ascribed to metal incorporation by solid solution or adsorption mechanisms into “mineral traps” (Jambor, 1994) as primary minerals (e.g. clay minerals), organic materials (amorphous colloids) and/or secondary minerals (e.g. jarosite, iron oxyhydroxides). The highest metal concentrations occur close to the wastes and rapidly decrease moving downstream some hundred of meters. The bulk analyses of the waste samples revealed high concentrations of heavy metals and typically low pH (Table 7). As and Pb are dominant in waste rock dumps (range: As = 55-1000 ppm; Pb = 30-27600 ppm) ; Zn and Pb in the flotation tailings (range: Zn = 150-9900 ppm; Pb = 450-4920); Cu and Bi in the roasting

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waste (range: Cu = 250-2280 ppm; Bi = 10-885 ppm). The investigated waste materials appear alterate (metal-depleted, acidified), and this could be ascribed to several factors: 1) primary mineralogical composition, and particularly the amount of metal sulphides and of pH-buffering phases (carbonates, chlorites and micas), 2) the age of mine wastes; 3) waste bodies morphology and grain size; 4) hydrological and chemical features of drainage waters.

The flotation impoundment contains tailings from the processing of pyrite ore during the period from 1957 to 1972. The impoundment is about 100x300m wide, with a maximum depth of about 10m, and dips gently to the north. There is a gradation from silt and sand in the south of the basin to mud and clay in the north. The northern zone is often flooded in the wet season and alteration is less pronounced in the north than in the south, presumably because oxidation of the waste is impeded by flooding (Benvenuti et al., 1997). Samples were taken at several sites, from the surface to depths of 40-70 cm. Sulphides and aluminosilicates are more abundant than at other localities, and the alteration sequence of the sulphides is apparently:

pyrrhotite>chalcopyrite=galena>arsenopyrite=sphalerite>pyrite.

The high susceptibility of chalcopyrite and galena to weathering is somewhat surprising. Jambor (1994) points out that the apparent resistance of chalcopyrite in several tailings impoundments may be due to its occurrence as locked inclusions in silicates and quartz. SEM/EDS analyses show that galena is usually corroded, with rare replacement by anglesite, and the scarcity of anglesite coatings may enhance the process of galena dissolution (Tanelli and Lattanzi, 1986). The comparative resistance of sphalerite to weathering may be due to a low Fe content; SEM/EDS analyses of sphalerite from tailings are consistent with microprobe analyses that report 1-9 mol% FeS (Tanelli and Lattanzi, 1986).

The uppermost sediments in the southern zone, down to about 20 cm depth, are sandy, with quite low pH (up to 5). Therefore, alteration of sulphides and alumosilicates is very advanced, and the main phases are quartz, iron-hoxyhydroxides, gypsum and minor jarosite.

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Claudio Bini

SoilsThe anomalous elemental concentrations in the solid fraction is not

restricted to the dump proximity. High contents of polluting elements (As, Bi, Cd, Cu, Ni, Pb, Z,n) in soils collected in the Merse river alluvial plain were recorded as far as 30 km moving downstream from waste areas.

Soils developed on, or proximal to, waste piles (“waste soils”) were compared to natural (“distal”) soils in the area. Generally, the soil evolution from waste is very limited, owing to the nature of parent material, the coarse grain size and the hydrological system. Waste soils show a shallow A-C profile (<60 cm), dark brown (7,5YR 3/2), subangular blocky structure; texture is coarse loamy. Natural soils are more deep (up to 100cm) and developed, with a marked discontinuity between topsoil and subsoil (A-2Bt-2C profile), where an illuvial (Bt horizon) formed. Colour ranges from dark brown (7,5YR 3/4) in the top, and dark reddish brown (5YR 3/4) in the bottom; structure varies from crumby to blocky, texture from loamy sand to clay loam; pH ranges from subacidic to subalkaline in the bottom, owing to the presence of calcareous fragments from the original limestone bedrock.

Compared with natural soils, waste-proximal soils (<50m) are very acidic, having a pH (2.5-4.8) lower than the former (3.5-7.5) and may include minerals from the ore bodies (pyrite, chalcopyrite), from ore processing (hematite), and from the weathering of these minerals (goethite, jarosite, alunite, copiapite, melanterite, anglesite and others). Distal soils (>50m from waste) are characterized by primary phases such as quartz, muscovite, chlorites, calcite, dolomite with minor amounts of kaolinite, illite, smectite rutile, ilmenite, zircon, monazite, derived from local bedrock weathering (phyllites, quartzites, limestone and flysch formation). Accordingly with the different mineralogy, the heavy-element content (especially Pb, Bi, As) in waste soils is appreciably higher than in natural soils (see Table 7), their average being higher than the concentrations considered dangerous or toxic for plants (Kabata-Pendias, 2004). Natural soils, either in close proximity to waste or downslope from them, are contaminated and acidified, almost in the topsoil. The contamination is probably caused by mechanical transport of primary and secondary minerals, including pyrite, hematite, sphalerite and jarosite, from the waste. Since only a part of the soil metal content is available for plants (Adriano, 2001), in order to evaluate the environmental (vegetation) hazard of the

44


investigated mine sites, bio-available (EDTA-extractable) elements were determined in soil samples, assuming extractable metal concentrations to be directly correlated with the amount of metals taken-up by plants. The amounts of EDTA-extractable metals are rather high (Cu = 0.3-22 ppm; Pb = 0.7-380 ppm; Zn = 0.1-860; Mn = 0.2-230; Fe = 4-1300 ppm), and exceed the limits usually considered toxic for plants (Kabata-Pendias and Pendias, 2001). In contrast, natural soils far from waste (>50m) contain only minerals such as quartz, muscovite, chlorite, calcite and dolomite, derived from the main lithologies outcropping in the area. Their metal content (total and EDTA-extracted) are below the limits of pollution. The soil metal contents commonly decrease with increasing depth, probably due to intense leaching and a longer residence time, and is likely controlled by morphology and grain size, that enhance water circulation. Yet, waste soils are typically water-unsaturated, and have a relatively high hydraulic conductivity. Aeolian transport and gravitational runoff, instead, are limited to the immediate vicinity of the mine sources.

In summary, several processes occur at the investigated sites in the Boccheggiano mine district:

• Sulphide oxidation and acid mine drainage production; • Decreasing acidification with distance from the mined area;• Metal content decrease with distance (dispersion halo);• Argillogenesis and clay migration;• Sulfatation and secondary minerals formation;• Runoff and leaching to adjacent streams, with consequent dilution.

4.1.5. Campiglia Marittima - Temperino Mine District

The Campiglia Marittima ore district has long been known for Cu-Pb-Zn skarn deposits enclosed within white marbles derived from contact metamorphism of Mesozoic limestone (Bertolani, 1958). These deposits lie 1-2 km E and NE of the Botro ai Marmi granitic stock (K/Ar age 5.7Ma), in strict spatial association with nearly coeval (4-5My) porphyry dikes (Corsini et al., 1980). Mining activity in the area dates back at least to Etruscan times (VII century BC), and flourished under the Romans (I century B.C.), in the Middle Age-Renaissance (X-XVI century A.D.), and in the XIX-XX centuries, until final closure in 1976. The mining district is

45

Claudio Bini

characterized by deposits of pyrite and mixed sulphides (Fe, Cu-Pb-Zn-Ag, Sb, Hg, Sn) and Au (Cipriani and Tanelli, 1983; Tanelli, 1985; Lattanzi et al., 1994 and references therein). In particular, two main styles of pyrite and polymetallic deposits have been identified (Tanelli, 1985): massive conformable bodies related to Palaeozoic-Triassic siliciclastic lithologies, and structurally controlled deposits associated with tectonic features of the Tertiary Apenninic orogeny or with Miocene-Pliocene magmatic rocks.

The morphology of the mined area is gently undulated, with elevation ranging from 150m to 450m a.s.l., warm Mediterranean climate (maT=16°C, maP=700mm). The vegetation climax is the Quercetum ilicis, and the present vegetation cover is a dense, deciduous forest (Mediterranean maquis) dominated by holm-oak. At some places, corresponding to mine spoil outcrops or to more exposed slopes, the forest is substituted by a shrubby formation (the so called garigue) dominated by rock-rose, or by discontinuous coverage with native grasses, especially fescue. Vegetation cover in the mineralized area is discontinuous, and, besides crusting lichens, hosts metal tolerant/accumulator plant species (Baker, 1981), as Cistus salvifolius, Inula viscosa, Silene paradoxa, Silene armeria, Festuca inops.

Waste rocks resulting from surface and underground mine working in the last two centuries were discharged in close proximity to the mine, and presently constitute a waste dump covering an area of about 0.1 km2

(Corsini et al., 1980; Baldini et al., 2001). The potential of the abandoned waste dumps to pollute the environment at these localities is enhanced to various degrees by the high topographic relief, the lack of vegetation cover and the proximity of the waste to streams.

The prevailing mineralogical phases in waste samples are quartz, ilvaite, hedembergite and pyrite, with smaller amounts of carbonates, Fe-Cu oxyhydroxides and chrysocolla. Until now, pH conditions (average value 6.4, range 5.7-7.0) have slowed the alteration processes of minerals, and favoured the absorption phenomena of leached metals onto oxyhydroxides surface.

A prolonged and continuous disposal of metals over time, however, may have important consequences for the environment and the plants life (Bini and Gaballo, 2006). For this reason, the mineralogy and geochemistry of both the waste material and the soils developed from, were investigated (Bini and Gaballo, 2006) in order to determine the extent

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of heavy metal dispersion in the conterminous land, and the related environmental hazard.

Three kinds of materials were sampled and analysed at selected sites in abandoned mine areas:

Waste-rock dumps Soils (50 profiles, both in mineralized areas and outside) Vegetation (selected plants, in spring and autumn)

Most dumps consist of coarse-grained waste rock from excavation; tailings from mineral processing also occur. The soils in the mineralized area, because of the generally steep morphology, are not very thick (mostly up to 40-50 cm, or up to 1m in terraced areas), neither very developed. They usually show coarse-grained textures with abundant lithic fragments, and are characterized by high permeability. Pedogenesis of waste dumps is normally minor, and mostly confined at the peripheral portions. Vegetation is herbaceous on the most recent mine wastes, while shrubby and arboreal plants colonize the older ones and the conterminous areas.

Waste-rock DumpsThe primary mineralogy is composed of pyrite and chalcopyrite,

ilvaite, hedembergite, sphalerite, Fe-oxides, quartz, muscovite and chlorite, with minor and variable amounts of secondary minerals (especially carbonates, clay minerals, jarosite and Fe-oxyhydroxides). Pyrite is commonly rimmed by Fe-oxyhydroxides, whereas dissolution features prevail in chalcopyrite. The particle size of the sulphides is variable (up to a few millimetres).

The prevailing mineralogical phases in waste samples are quartz, ilvaite, hedenbergite and pyrite, with smaller amounts of carbonates (calcite, dolomite and smithsonite), Fe-Cu oxyhydroxides and chrysocolla.

Waste dump materials contain relatively high amounts of toxic metals (average values, in wt%: Cu= 1.3, Pb= 0.2, Zn=1, As=0.01, Bi=0.02) exceeding the maximum permitted limits according to Italian legislation (D.M. 152/2006). A relevant part of these metals may be transferred to conterminous soils by chemical alteration, runoff and wind transport, thus determining a potential concern to the environment.

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SoilsThe soils of the mineralized areas show a noteworthy spatial

variability, as evidenced by a different degree of evolution. Entisols (Lithic, Typic, Spolic Xerorthents) are common on recent mine dumps (<100y), while Inceptisols (Haploxerepts and Dystroxerepts) and Alfisols (Haploxeralfs and Rhodoxeralfs) are frequent on more ancient dumps or in the conterminous areas, where a lithological discontinuity occurs between the “primary” parent material (limestone, shale, metamorphic or magmatic rocks) and the “secondary” mine waste (polycyclic soils, chronosequences). Immature Entisols are characterized by a thin solum (<30cm), little organic matter accumulation (mean 14g/kg organic carbon, range 1-33 g/kg), dark brown (10YR3/3) to reddish (5YR4/6) colour, coarse texture (sandy loam to loamy), and subalkaline pH (mean 7.4, range 6.9-7.8). An environmental consequence of the high content of toxic heavy metals (see Table 8 and Figure 8), in combination with reduced soil thickness, is a discontinuous vegetation coverage.

The soils developed from old mine dumps, or in the proximal parts (<0.5 km) of the dumps, are characterized by a solum >50 cm thick, sandy loam to loam texture, blocky structure, slightly acidic pH (mean value 6.3, range 4.9-7.7), humus accumulation (up to 14% organic matter in the A horizon), moderate to low cation exchange capacity (mean 20 cmol(+)/kg), with significant desaturation (base saturation <60%). Generally, they have distinct A-B-C horizonation and a well formed cambic horizon. Therefore, they are Inceptisols (Spolic Haploxerepts or Spolic Dystroxerepts, see Table 8 and Figure 9). Data (not reported) indicate relevant differences and a remarkable polycyclic evolution, owing to the superposition of mine spoil over the normal soil. Colour, texture, reaction, and cation exchange capacity are the most prominent features that present major differences. Soil horizons show dark brown (7.5YR3/2) to dark reddish brown (5YR 3/3) colour, well individualized structure, from crumby to fine blocky peds. Texture is coarse (sandy loam to loam) in surface horizons sampled on mine spoils and loamy to clayey underneath. Soil reaction is slightly acidic (pH 6.3) at the surface, subalkaline (pH 7.4) and base-saturated at the bottom. Cation exchange capacity increases with depth, from 15 to 25 cmol(+)/kg.

Soils described and sampled at major distance (>0.5km) from the mine dumps present little evidence of mine spoil in the profile. Sulphide

48


minerals are found especially at the surface, as revealed by mineralogical and chemical composition (see Table 8 and Figure 10). An abrupt textural change indicates a marked discontinuity between the upper and lower part of the soil profile. The upper part (A and E horizons) has dark brown (10YR2/2) to yellowish brown (7.5YR3/4) colours, loam to sandy loam texture, crumby structure, high organic carbon content (mean 35 g/kg, range 21-57), and subalkaline pH. The lower part (Bt horizon) presents reddish colours (5YR3/4 to 2.5YR3/4), clay content increase, with clay loam to clay texture, organic carbon and pH decrease (mean 6.4, range 5.1-7.9), and carbonate is absent. These features are consistent with soil development from limestone in the Mediterranean environment (the so called “Terra rossa”). Therefore, they are classified as Alfisols (Spolic Rhodoxeralfs or Spolic Haploxeralfs) or, alternatively, as Spolic Xerorthent over Typic Rhodoxeralf (or Haploxeralf).

Soil Mineralogy and GeochemistryWaste soils are characterized by high contents of primary and

secondary metalliferous phases (sulphides, skarn-silicates, oxyhydroxides, carbonates, sulphates, Fe-Cu oxyhydroxides). Phyllosilicates are present in limited amounts in pedogenic horizons. Chlorite and mica may form from skarn silicates and mine spoil as well, while illite is likely to be inherited from the terra rossa. Moreover, aeolian dust could have contributed to soil mineralogy, especially quartz and phyllosilicates, as reported by Bini et al. (2006) for similar soils.

The bulk chemical composition of waste soils indicates (Table 8) high amounts (up to 73%) of silica and alumina (15%) in A horizons of Entisols. Total Fe, Mg and Mn, instead, present higher amounts in C horizon (14.70%, 3.36%, and 4.36%, respectively) in comparison to the A horizons. Sodium and K decrease with depth, while Ca increases (6.47%). Titanium concentration is rather constant, as expected for a very stable element. Heavy metal concentrations are higher in the C horizon than at the surface, being related to the original composition of the spoil material.

Lesser amounts of Si, Al, Mg in surface horizons, in comparison to subsurface horizons, are recorded in Inceptisol (see Table 8). Instead, Fe, Mn, and Ca occur in higher concentrations at the surface than at the bottom. Titanium is rather constant, as in the previous Entisol. Extremely high concentrations of heavy metals, contributed by spoil material, are

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Claudio Bini

recorded in the whole profile; their distribution with depth corresponds to soil discontinuities.

Mine soils over Alfisols present (see Table 8) a general increase in Si and Al, and a decrease in Ca and Mg, with depth, while Fe, Mn, K, and Ti are quite constant. Heavy metal concentrations are high, and are distributed irregularly, as a consequence of differential contribution from spoil parent material.

As already mentioned, most soil pH values fall within one unity from neutrality (range 4.9-7.4, average value 6.4). Until now, these pH conditions have slowed the alteration processes of minerals (as confirmed by minor alteration of sulphidic phases), and favoured the adsorption phenomena of leached metals onto oxyhydroxides surfaces. Therefore, the solid phase of waste soils is strongly enriched in metals, whose levels overcome the target values of current Italian legislation (D.M.152/2006). Their concentration, however, depends on the distance from mining areas and on the age of waste.

The results of leaching tests (Baldini et al., 2001) indicate a low degree of exchangeability (i.e., bioavailability) of all the metals, with Zn>Fe>Cu=Mn. (average values, in % of total metal contents: Zn=3; Fe=0.5; Cu=Mn=0.3). Since the bioavailable metal fraction is quite low, phytotoxicity is quite unlikely; yet, it is noteworthy to consider that the absolute metal content of soils cannot lead to progressive metal enrichment in plants. Nevertheless, a prolonged and continuous disposal of metals on the land may have important consequences for plants’ life as well.

4.3. Apuane Alps

A set of papers (Mascaro et al., 1999; 2000; Benvenuti et al., 1999; 2000; 2001) has been addressed to evaluate the environmental impact determined by past mine activity and metal production at the Bottino mine, Apuane Alps. The evaluation need is corroborated by the fact that the investigated area is located within the Natural Park of Apuane Alps, a protected area with high natural interest.

50

Table 8. Bulk chemical composition of selected soil profiles (<2 mm fraction; major element concentrations are expressed in weight %, trace elements in mg/kg)

Soil Profile (SSS, 1999)

Soil Horizon

Na2O MgO Al2O3 SiO2 P2O5 K2O CaO TiO2 MnO Fe2O3 Cu Pb Zn

Spolic A1 1.10 1.45 12.53 72.64 0.06 2.07 0.63 0.63 0.19 6.66 170 78 260Xerorthent A2 0.90 1.66 15.23 66.35 0.06 2.39 0.62 0.69 0.24 8.15 260 24 120Waste soil 2C 0.43 3.36 12.70 41.54 0.40 1.70 6.47 0.73 4.36 14.70 1000 840 2000Spolic A1 0.49 2.04 6.60 43.57 0.05 1.23 3.32 0.21 3.91 33.81 9665 9200 15000Dystroxerept A2 0.18 1.39 3.87 47.35 0.04 0.69 4.55 0.12 3.10 44.46 14100 4100 5840

Bw 0.62 2.99 10.52 37.53 0.07 1.76 1.73 0.41 3.16 23.78 3790 11900 19800Proximal soil 2Bwb 0.19 1.46 13.72 37.17 0.03 0.66 6.22 0.21 3.92 41.33 14900 3770 6240

2Bwb’ 0.21 3.09 13.75 64.41 0.08 2.54 4.73 0.54 1.30 10.61 7310 1100 26403BCb 0.18 3.73 16.62 65.94 0.04 2.68 0.50 0.56 0.89 6.38 3700 340 24003Cb 0.18 7.35 22.16 49.14 0.05 3.04 0.69 0.70 2.83 9.61 9230 320 4300

Spolic A 0.02 4.26 15.15 45.95 0.16 3.05 4.75 0.79 1.50 10.38 420 420 920Xerorthent E 0.21 4.55 15.72 60.29 0.06 2.84 0.89 0.73 1.03 9.83 280 460 620Over EB 0.22 3.86 16.42 58.84 0.07 2.95 0.90 0.71 1.18 10.04 590 1500 1260Rhodoxeralf 2Bt1 0.35 2.69 18.47 60.35 0.06 3.34 0.65 0.81 0.95 9.21 390 470 780Distal soil 2Bt2 0.31 2.45 24.33 52.82 0.05 3.21 0.64 0.77 1.04 10.91 140 330 510

3C 0.05 11.52 22.67 45.02 0.04 3.72 0.89 1.00 0.57 9.98 16 65 540

Table 9. Average and range of trace elements concentrations, mean values of pH and texture at different sites in the Bottino mine area. (Fe, Mn, S are expressed as %; As, Cd, Cu, Pb, Sb, Zn as mgkg-1).

Fe Mn Pb Zn S As Cd Cu Sb pH texture

Mine waste (1593)

9.8 (6.9-12.5)

0.6 (0.3-0.9)

0.5 (0.05-1

0.3 (0.3-0.4)

0.7 (0.5-0.8)

86 (3-174)

30 (24-38)

442 (272-568)

110 (3-170)

6.8 (6.5-7.3)

Gravel, pebble

Mine waste (1840)

7.6 (4.1-10.4)

0.4 (0.05-0.7)

0.6 (0.06-3.9)

0.6 (0.01-3.2)

1.0 0.2-4.9)

75 (3-150)

75 (5-485)

177 (40-430)

82 (18-260)

6.8 (3.5-7.9)

Gravel, pebble

Mine waste (1929)

7.8 (6.8-9.5)

0.3 (0.1-0.6)

0.8 (0.2-2.0)

0.7 (0.1-2.5)

1.0 (0.4-2.7)

266 (95-576)

45 (5-140)

384 (133-983)

224 (41-680)

6.2 (4.7-6.6)

Gravel, pebble

Mine soil (1995)

5.7 (7.2-8.7)

0.3 (0.3-0.4)

0.7 (0.2-2.4)

0.9 (0.3-1.3)

1.5 (0.4-3.1)

470 (100-760)

60 (24-90)

188 (70-320)

700 (110-1750)

5.7 (5.3-6.1)

Sandy skeletal


5.6 (5.0-6.6)

0.1 (0.07-0.1)

0.05 (0.02-0.1)

0.2 (0.1-0.2

0.04 (0.01-0.2)

112 (90-130)

6 (5-9)

52 (35-70)

16 (7-28)

5.1 (4.9-5.3)

Sandy loam

Distal soil (1995)

4.9 (3.0-5.8)

0.02 (0.01-0.03)

0.03 (0.02-0.03)

0.2 (0.2-0.3)

0.03 (0.01-0.03)

70 (50-80)

7 (5-11)

23 (6-33)

7 (5-8)

4.6(4.0-5.0)

Loamy sand

Adapted after Mascaro et al. (2000; 2001b).

Claudio Bini

The ore deposits in the Apuane Alps district, northern Tuscany, has been exploited more recently than in Southern Tuscany. The Pb(Zn)-Ag vein deposit at Bottino was mined in the 16th and 19th centuries, with recorded production totalling 4000 t Pb, 600 t Zn and 22 t Ag. The quartz-carbonate-sulphide veins are hosted by metamorphosed phyllites and acidic volcanites of the Palaeozoic basement (Benvenuti et al., 1999). The ore bodies lie typically next to the Bottino creek, over steep slopes (about 40°) at high elevation (500-700m asl). The vegetation cover is discontinuous and is composed of mixed forest (prevalently oak); terraced slopes cultivated with chestnut occur at distal sites. Climate is humid temperate (maT=14°C, maP=1800mm). Soil development over the dumps is moderate (<40cm), and mostly confined at the borders. Both proximal and distal soils present more developed horizonation, with total thickness >100cm.

Waste piles in the area include excavation waste and tailings from hand-picking and jigging. Five dumps were investigated; they are elongated heaps that extend some tens of metres downslope, with widths of about 10 m and a maximum thickness of 2-4 m. The material is coarse grained (pebbles and sand).

The moderate thickness, the coarse-particle size and the generally steep slopes suggest that most, if not all, waste bodies are within the vadose zone (Benvenuti et al., 2000). As very little action was undertaken to minimize the environmental impact during and after exploitation, many waste piles are left scattered over the land, and drainage from mining tunnels freely runs into the stream network. Waste bodies cover an area of approximately 5300 square meters, with thickness of 2-5 m. The total volume of waste material is estimated between 20.000 and 50.000 cubic meters.

The mineralogy and chemistry of the analysed surface material are highly heterogeneous, and heavy-metal concentrations are typically high. Weathering processes are more pronounced in both the older and finer material than in the more recent and coarser one.

Primary minerals (i.e. minerals originally present in the mineralization and/or in the country rocks, irrespective of their origin; see Jambor, 1994 for definition) are much more abundant than secondary minerals. Among primary mineralogical phases, quartz and white mica are ubiquitous, and quite abundant. Other common silicates include chlorite, albite, and

53

Claudio Bini

tourmaline. Carbonates are mainly represented by terms of the siderite-magnesite and dolomite-ankerite solid solutions; calcite is comparatively rare. The most abundant sulphides are sphalerite, galena, and pyrite; chalcopyrite, pyrrhotite, arsenopyrite, Ag-bearing tetrahedrite and boulangerite also occur. Clay minerals occur as primary (illite and chlorite) and as secondary phases (kaolinte, montmorillonite, vermiculite).

The main secondary minerals (i.e. secondary phases that developed in situ by supergene processes) are goethite, lepidocrocite, pyrolusite and cerussite. Yellow rusty crusts are present in some samples. Other phases occurring in small amounts that preclude conclusive identification, include poorly crystalline or amorphous Fe-Mn-Al-hydroxides, containing in places appreciable amounts of other metals (up to 20%Zn, 25% Pb, 7% Cu, 2% Ni, !% Co). As suggested by Benvenuti et al. (2000, 2001) and by Mascaro et al. (2000), these may result from:

a) submicroscopic admixtures of separate phases; b) isomorphogenous substitution in the lattice; c) surface adsorption;d) any combination of the three possibilities.

Minor minerals are ferrihydrite, pyrochroite, cuprite, malachite, Fe-sulfates, unidentified Fe and Pb sulfarsenate, amorphous or cryptocrystalline material, and native gold. The gold is closely associated with Fe and Mn-oxyhydroxides and its texture suggests that it is the result of secondary redistribution (Porto and Hale, 1995). A comparison of mineralogical data and the pH measured in the waste material indicates that the alteration occurs under nearly neutral pH conditions. Yet, most pH values fall within one unity from neutrality (see Table 9).

The relative alterability of sulphides is consistent with what is commonly observed in sulphide mine waste elsewhere (Jambor, 1994), and follows the sequence:

pyrrhotite > galena-sphalerite>chalcopyrite-arsenopyrite-pyrite.

Oxidation of pyrrhotite generates Fe-oxyhydroxides, and galena forms cerussite, while sphalerite presents more pronounced dissolution, accompanied by rims of Fe(Mn) oxyhydroxides (Benvenuti et al., 2000).

54


The main reactions controlling the generation of acidity, which in turn promotes further metal release from primary sulphides, are oxidation of pyrite and dissolution of carbonates.

The potential threat to the environment represented by the high metal contents of this mine waste is mitigated by their entrapment in comparatively stable phases such as goethite, lepidocrocite, pyrolusite, and cerussite, whereas entrapment in metastable or easily soluble phases such as ferrihydrite and Fe-sulphates is obviously ephemeral.

In the Bottino mining district, the abandoned waste piles contain variable amounts of heavy metals that are of potential concern for the environment. The observed neutral pH values favour a relatively efficient fixation of heavy metals in stable phases (e.g. Fe-Mn oxyhydroxides and cerussite), in such conditions. Supergene alteration of primary minerals in the dumps may occur in two ways:

in situ pseudomorphic replacement of secondary phases (e.g. galena/cerussite):

PbS + CO32+ + 2O2 → PbCO3↓ + SO4

2-, and/or leaching and dissolution. The elements that are released may

eventually reprecipitate as secondary phases (e.g. galena/anglesite):

PbS + 2O2 → Pb2+ + SO42- = PbSO4↓

The most relevant results of the mineral chemistry studied by SEM/EDS semiquantitative analysis (Benvenuti et al., 2000) are:

Fe and Mn-oxyhydroxides contain appreciable amounts of Co, Cu, Ni, Sb and S;

Cerussite may contain traces of Zn; Some Fe-sulphates contain variable amounts of Ag, Cu, Pb, Sn,

and Zn; Fe-Pb sulfarsenates do not contain other metals; The amorphous and cryptocrystalline substances contain

appreciable amounts of Cu, Pb, Sb, and Zn.

The presence of large amounts of secondary minerals suggests that the waste materials are slowly approaching a mature stage of supergene

55

Claudio Bini

alteration. However, given the low amounts of both sulphides and calcite in the area, and the nearly-neutral pH, acidification processes are quite limited, and therefore the alteration phenomena occurred, until now, with no significant environmental pollution.

The metal contents of soils developed from waste material, and those proximal and distal from the mine dumps, have been investigated by Mascaro et al. (2000; 2001b), together with wild plants growing at the same sites. The results concerning waste and soils are summarized in Table 9.

Mine waste typically are enriched in metals with respect to waste soils. Metal contents generally decrease with the age of the mine spoil, and with distance from the metal source, as observed also by Bini and Gaballo (2006). Yet, soil evolution with time, and the distance from the mine spoil, greatly contributes to the dilution effect: pedogenic processes such as acidification, humification, leaching, mobilize metals in different forms (soluble, chelate, or adsorbed), so that they may be leached away from the soil system to groundwaters.

Waste soils (Lithic Udorthents in the USDA soil taxonomy, 1999) show little thickness (<50cm), coarse-grained texture and abundant lithic fragments, with high permeability; profile evolution is limited, with A-C horizonation, and their characteristics reflect those of the parent material. Proximal and distal soils, instead, are more developed, with clear ABwC horizonation (i.e. they are Inceptisols: Dystric Eutrudepts or Umbric Dystrudepts in the USDA soil taxonomy, 1999). Colour change, from gray in entisols to reddish brown in inceptisols, is the most evident character; texture is finer in inceptisols than in entisols; organic matter content is quite low (<1 % O.C.), and scarcely humified; the most common humus feature is the moder with subacid reaction. The acidic pH could be the main cause of metal release from the solid phase. However, the metal content in soils is likely related mainly to mechanical transport (runoff, aeolian, gravitative) of metal-bearing material from mine waste, than to in situ transformation of primary phases.

Heavy metal and sulphur average contents in both waste and proximal soils usually exceed the maximum permitted values for farming soils of the Italian legislation (DM 152/2006); distal soils, instead, present metal contents below such limits, except for As. The almost acidic nature of soils is likely due to the parent material and to the presence of organic acids

56


produced by the forest ecosystem, and the scarcity of acidity buffers such as calcite and dolomite. Extraction tests show that the metal contents of exchangeable fraction do not exceed 1-3% of the total concentrations (Mascaro et al., 2001b); on the contrary, in agreement with mineralogical data, metals bound to sulphide (± organic matter) fraction is comparatively high, ranging from 20% (Mn) to 80% (Pb). The carbonate-bound fraction is commonly 10-20% of total, except for Mn (54%), probably “trapped” as carbonate. The percentage of Mn, Fe and Pb in the reducible fraction (i.e. that due to the dissolution of Fe-Mn-(Pb) oxyhydroxides) are nearly 10% of the total metal concentrations; lower percentages have been calculated for Cu and Zn (1-7%). The relative high percentages of Fe, Zn and Cu in the residual fraction (25-35%) may be explained by the occurrence of chlorite and sulphides encapsulated within quartz grains.

The occurrence of high metal contents in some plants can be accounted for by continuous, although limited, supply of metals from the contaminated material. The reduced thickness, coarse particle size and high metal content of mine waste hinders the growth of arboreal vegetation at the top of waste piles. The same factors are also responsible for the reduced pedogenesis of the waste material.

4.4. SARDINIA

Sardinia, the biggest Italian island, for millennia has been the main metalliferous region of Italy. Historical mining was carried out for long time in various mining districts, in absence of effective regulations and controls on the environmental impact, and was characterized by disregard of environmental issues, with consequent diffuse contamination (Da Pelo et al., 2009). Only in the last decades, Regional Authorities decided to counteract the environmental impact of abandoned mines, and to restore metal-contaminated areas. Since that time, numerous natural areas and mine-archaeological parks have been realized in the island, where the cessation of mining activity left large quantities of mine wastes on dumps and flotation tailings, estimated at about 45 Millions m3 for the whole mining district (Cidu et al., 2009). Several studies, therefore, have been carried out in the last decade on these abandoned mine areas (Caboi et al., 1999; Cidu and Fanfani, 2002; Frau and Ardau, 2003; Musu et al., 2007;

57

Claudio Bini

Trois et al., 2007; Cidu et al., 2009; Da Pelo et al., 2009; Frau et al., 2009), with the aim of understanding the environmental impact of mining operations.

A century of exploitation of huge galena-sphalerite deposits hosted in the Iglesiente-Sulcis mining district, for Pb and As production (Frau et al., 2009), has caused remarkable environmental impact, mainly due to discharging tailings from the flotation basins directly into streams that drain the mined areas. This has caused diffuse contamination in the whole catchment, and the dispersion of highly contaminated materials over a distance of about 10 km downstream from the mine. The geology of the district consists of ore bodies embedded in Cambrian limestones and dolostones with gently undulating morphology. Climax vegetation is the Mediterranean maquis, here in a degraded stage, with a discontinuous cover composed mainly of herbaceous and shrubby vegetation. Climate is typically Mediterranean, with warm and dry summer and mild humid winter (maT = 16°C, maP = 650mm). Soils are scarcely developed Entisols (USDA soil taxonomy) mixed with bedrock outcrops. Some red soils too outcrop in the conterminous areas not affected by mine waste.

An inventory of the mineral resources of the Southern Sardinia mine district has been made recently by Cidu et al. (2009) and Frau et al. (2009), who collected samples from waste-rock dumps, flotation tailings and stream sediments.

The main mineralogical composition of the waste dumps consists of quartz, K-feldspar, chlorite, muscovite and biotite, derived from the bedrock. In mine dumps, secondary phases are the abundant presence of anglesite as an oxidation product of galena, cerussite (linked to the carbonatic fraction) and resistant minerals such as Pb-Fe sulphate (plumbojarosite). Tailings samples contain minor amounts of montmorillonite and gypsum, and traces of oxyhydroxides.

According to Frau et al. (2009), surface waters, with the exception of two acidic samples, are neutral or slightly alkaline (pH 7-8) and oxidizing (Eh 0.4-0.6 V). These chemical-physical conditions enhance dissolution of sulphates deriving directly (anglesite), or indirectly (gypsum), from sulphides oxidation; this determines metal release at sites close to the flotation basin, and a metal concentration decrease about 1.5 km upstream. Metal attenuation in surface waters does not depend on mixing, but rather on a removal process, following the sequence Pb>Cu>Zn>Cd suggested by

58


Caboi et al. (1999) for another mine area. Another possible removal mechanism might be represented by precipitation of secondary plumbojarosite as small grains dispersed in a ferrihydrite mass or as coatings. Indeed, most of the plumbojarosite contained in stream sediments/tailings was formed by direct precipitation from the flotation basin. SEM/EDX observations carried out by Frau et al. (2009) on waste-rock dumps actually show that galena crystals have an alteration rim composed of anglesite, while ferrihydrite and Pb-jarosite coatings formed on quartz grains in stream-bed sediments and flotation tailings.

In the recently open gold deposit at Furtei, exploitation was preceded and accompanied by studies on the environmental impact. Da Pelo et al. (2009) collected solid samples, including mineralized rocks and related proximal soils, weakly mineralized rock (and distal soils), and tailings from impoundment. Moreover, water collected immediately after a heavy rain event was assumed to represent the natural leaching of exposed materials.

The total concentration of selected elements in the solid phases is reported in Table 10. Iron is mostly abundant in mine-waste (average 5.9%, range 3.2-7.8), decreasing down to 1.7% in tailings and to 0.3% in distal soils, as well as Mn. Consistently, also trace elements (As, Cd, Cu, Pb, Zn) present higher concentrations in waste-rocks than in tailings and soils, as expected as a consequence of lower pH, minor grain size and a dilution effect determined by distance from the mineral source.

Table 10. Average trace elements concentrations, mean values of pH and texture at different sites in the South Sardinia mine area

Fe Mn As Cd Cu Pb Zn pH texture

Mine waste

5.9 1.20 225 0.6 982 68 107 3.3 Gravel, pebble

Tailings

1.7 0.75 280 <0.5

500 94 41 6.7 Coarse Loamy

Distal soils

0.4 0.33 32 <0.5

35 87 56 5.9 Sandy loam

Fe, Mn are expressed as %; As, Cd, Cu, Pb, Zn as mgkg -1. Data from Trois et al., 2007 ; Da Pelo et al., 2009; Frau et al., 2009.

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Claudio Bini

Leaching experiments show (Frau et al., 2009) intermediate to low concentrations of minor components, with toxic elements mostly immobile, being related to the residual fraction, in a near-neutral pH. Exception to this general behaviour is given by As, whose relatively high, albeit variable, percentage of extractability is linked to the large availability of pyrite. Yet, pyrite is abundant in mineralized rocks, and little in distal samples. Tailings contain minor amounts of montmorillonite and gypsum, along with traces of Fe-oxyhydroxides and pyrite.

High amounts of S, Fe (from pyrite) and metals (especially Cu, As and Ba) are present in dumps.

Pyrite oxidation in acidic environment (average pH 3.3; range 2.6-7.6) promotes the oxidation of other sulphides, the release of metals and the formation of soluble secondary minerals. The majority of base metal sulphides are solubilised in an abiotic manner in acidic conditions by sulphuric acid or by an oxidising agent such as ferric ion. The solubilisation kinetics may be increased by microorganisms such as bacteria to such an extent (10 5 times) that they can be regarded as producers of sulphuric acid. Microbiological viability tests (Trois et al., 2007) were aimed at recognizing the presence of acidophilic chemolithoautotrophs in the dump layers. The oxidation and solubilisation of mineral metal sulphides, catalyzed by chemolithoautotrophic acidophiles of the genus Acidithiobacillus ferrooxidans, are the main causes of acidic rock drainage. It was found that they were absent from the upper layers, whereas in the underlying layers they are pronouncedly present, due to an increase in moisture content. Yet, there is evidence of a strong dependence between the presence of a acidophilic chemolithoautotroph microflora inside the dump and the quality of effluents (Trois et al., 2007). Moreover, the presence of bacteria and/or of dissolved ferric iron from pyrite oxidation speeds up mineral dissolution, with consequent release of As and Cu to the water. A comparatively slow reaction rate can still result in the release of a harmful amount of contaminants.

Typical features of the process are:

the pronounced acidity of the effluents after irrigation;

60


the repetitive presence of metal ions concentration peaks on resumption of each irrigation, regardless the extent of dry periods;

the almost complete and lasting disappearance of metal ions from the effluent after neutralization of the inflowing water, coupled with a drastic reduction of the viable microflora.

The highest potential threat for the environment is mostly represented by mineralized rocks exposed in waste dumps and open pits. The waste dumps associated with dismissed mining activities produce, during rainfall events, acidic solutions containing potentially toxic elements (As, Cd, Co, Cu, Ni, Pb, Zn) in concentrations that exceed the discharge limits. After a sufficiently long rainy period, the effluents are low in toxic metals, and this may generate the impression that the contamination potential of dumps can be reduced by sufficiently long irrigation strategies (Trois et al., 2007). Moreover, mechanical removal (wind blown or gravitational) and/or rain leaching may contaminate conterminous farming soils and groundwaters (Cidu et al., 2009). Therefore, mitigation actions must be addressed to land reclamation. A natural attenuation of acidity and metal load just occurs upon interaction between minerals and waters in the vadose zone, as ascertained by ongoing monitoring programs (Da Pelo et al., 2009). Tailings confinement to a restricted site should minimize their environmental impact. The remediation of the environmental impact posed by dumps of rocks containing base metals has been successful in a number of mine sites, taking into consideration the rock geochemistry, the dumps geotechnical properties and the climatic conditions. Revegetation of exposed rocks is ongoing at some sites, based on the presence of an active microflora (Trois et al., 2007).

4.5. VENETIAN TERRITORY

The Venetian territory has been inhabited since long time, and was colonized intensively by Celtic and Roman people. Ore exploitation, however, took place especially during the late Middle Age and the Renaissance, when the Venice Republic dominated the most part of North-Eastern territories, where forests, for building houses and ships, and metals, smelted for the coinage and the armaments, were easily available.

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Claudio Bini

In the territory there is a number of small ancient mines (Gares, Forno di Canale, Vallalta, Zoldo), located in mountain areas of difficult access, with the exception of the one in the Imperina creek valley, all closed since the end of XIX century.

4.5.1. Valle Imperina

Valle Imperina, a Cu-Fe-Zn-(Pb) mixed sulphide ore deposit, was the most important mine site in the Venetian territory since 1400, and during five centuries, until final closure in 1962, supplied copper to the Venice Republic for coinage and armaments. Ore exploitation was clearly expanding since the opening of the mine: metallic copper production was 15 tonnes in 1574, 62 tonnes in 1669, and 120tonnes in 1788, equivalent to approximately half of the necessity of the Republic, and about 200 tonnes in 19th century. Afterwards, since the 1868, the decline of international price determined the abandonment of exploitation for copper, and pyrite ore was exploited for sulphuric acid production, until final closure.

The Imperina creek valley is located in the mountain district of Belluno (North-East Italy), with an altitude ranging between 543m a.s.l. and 990m a.s.l., and oriented in the SW-NE direction. The geological substrate consists of dolomite rocks (Upper Triassic) on the right side and the predominantly metamorphic basement (Pre-Permian) on the left side, while at the bottom the calcareous-arenaceous complex of Werfen (Upper Permian - Lower Triassic) is outcropping (Bini et al., 2010; Fontana et al., 2010). Even if no human settlement could be found presently in the area, many buildings and tunnel outlets still bear witness to the past mining activity (Figure 12). Part of the area (right side and a portion of the bottom) lies within the National Park of the Belluno Dolomites. The Imperina stream crosses the valley, along a tectonic contact between the metamorphic basement and the Mesozoic dolomite rocks. The mineralized area of Valle Imperina, which is located along the above contact, is a deposit of mixed sulphides, composed primarily of cupriferous pyrite, pyrite and chalcopyrite, with minor amounts of other metallic minerals. Its exploitation has continued almost uninterruptedly from the XV century until the year 1962; copper and sulphur were the main products extracted. Until the beginning of the XX century, copper was extracted and processed

62


directly in situ through roasting, a method with a severe impact on the area due to acid rains formation and intensive wood cutting. Yet, in early ‘900 vegetation in the area was lacking, due to cuttings and acid rains derived from SO2 production after metal roasting; presently, the whole area is naturally vegetated with mixed forest. Climate is humid temperate, with maT = 13°C and maP = 1250mm. A preliminary soil survey was carried out in the area in the last years by Bini et al. (2004) and Bini and Zilioli (2010), with the aim of characterizing soil genesis and evolution in the alpine environment. More than 70 soil profiles were described and analyzed; of these, eight profiles were selected for specific environmental analyses by Fontana et al. (2010). In general, the soils encountered in the survey are shallow (30 cm to a maximum of 100 cm) and undeveloped, with little presence of diagnostic subsurface horizons, and this applies particularly to those located in the areas affected by mining and metal processing, while those sampled as control are more developed. The pH varies from about 4.0 to nearly 8.0, depending on the nature of the substrate; the highly acidic pH values found in some soils are probably due to the alteration process of iron sulphides (pyrite and chalcopyrite) in the soil and substrate (Delgado et al., 2009). The texture is typically loamy, sandy-loamy or silty-loamy. The structure is usually weak and in some cases soils tend to be structureless and loose. The cation exchange capacity is low for all the soil samples, except for the profile on dolomite. The abundance of the soil skeleton is variable; in some profiles (1 and 4) it consists of waste from processing of ferrous minerals and coal from roasting, which show a clear anthropogenic influence.

The concentrations of heavy metals in soil samples are shown in Table 11. Comparing the values found with those of control levels, according to the Italian legislation (D.L. 152/2006) and the world averages (Angelone and Bini, (1992), the area seems not contaminated with Ni, Cr and Mn, while there is a contamination by Zn and a high contamination by Cu, Pb and Fe, which are present in high concentrations, particularly in the soil nearby the areas affected by mining and ore processing (profiles 1, 2, 4, 6).

Soils sampled in the immediate vicinity of the stream Imperina (profiles 3, 5) are less polluted than others in the valley, even though they are included in the area of greatest influence of human activities. This is probably due to the leaching of water and the establishment of periodic reducing conditions that increase the mobility of most elements considered,

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favouring the removal from soil and amplifying the risk of water contamination (Adriano, 2001).

Figure 12. The old mining buildings in the Imperina Creek Valley in early 1900. Note the lack of vegetation in back mountain. Presently the buildings are restored, as well as the vegetation cover, and host a small museum and an hostel. (Photo Archive Mining Technical Institute, Agordo)

The distribution of selected heavy metals (Cu, Fe, Pb and Zn) along the soil profile shows a general tendency to metal accumulation at surface. This is particularly true of Pb, and is consistent with pedogenetic processes occurring in the area. Significant variations of metal concentration with depth mark some discontinuities recorded in the profile morphology as well.

Table 11. Concentration of metals in soils of Imperina Valley, average values of reference (Angelone and Bini, 1992) and limit values in the

Italian legislation (D.M. 152/2006)

Ni Cr Cu Pb Zn Fe Mnmg mg kg-1 mg kg-1 mg kg-1 mg kg-1 % mg kg-1

64


kg-1

Prof. 1 waste soil average

< DL < DL 3159.49 23605.52 1588.92 52.31 256.07

range - - 2198.19-4063.93

20814.97-28154.16

1162.91-1786.37

50.87-53.55

169.81-440.46


< DL 41.07 2494.35 7057.91 980.02 17.90 506.38

range - 20.49-59.76

1936.30-3367.14

2497.47-14634.89

799.73-1188.32

9.66-31.86

279.99-669.44

Pr. 3 stream bed average

54.77 95.31 1122.31 372.98 734.45 6.85 1075.96

range 49.12-60.42

88.85-101.77

526.44-1718.18

227.77-518.18

471.99-996.90

6.18-7.52

985.68-1166.24


< DL 39.54 3093.55 5815.92 1192.20 40.30 84.09

range - 24.26-48.28

1768.54-4419.89

1435.79-14619.29

271.73-2423.03

19.72-58.22

31.83-174.59

Pr. 5 stream bed average

48.64 95.75 512.02 512.02 476.66 5.60 1199.23

range 46.62-50.65

98.71-92.80

524.29-499.75

205.46-293.75

430.85-522.47

5.91-5.29

1139.51-1258.96


< DL < DL 1639.04 8256.22 1338.37 43.23 215.32

range - - 502.51-2333.92

397.04-12026.83

408.52-2566.27

4.46-56.82

114.77-522.42

Prof. 7 distal soil average

27.72 163.63 53.98 64.08 71.23 3.68 811.13

range 16.91-57.68

147.10-177.17

30.79-97.91

52.14-72.43

103.14-40.80

3.14-4.02

582.82-1024.67

Prof. 8 distal soil average

13.66 31.37 1224.44 570.79 816.79 1.48 181.72

range 12.18-16.24

26.85-34.17

282.76-1774.73

342.74-762.99

576.80-1039.90

1.28-1.69

151.91-202.07

Italian average (1992)

46 100 51 21 89 3.70 900

International 40 200 20 10 50 - 850

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Claudio Bini

average (1992)Residential Limits(DL 152/2006)

120 150 120 100 150 - -

<DL = less than the detection limit.

Table 12. Linear correlation coefficient calculated on the concentrations of metals in soils in all horizons

Ni Cr Cu Pb Zn Fe MnNi 1.00 0.31 -0.13 -0.41 -0.09 0.80 0.86Cr 0.31 1.00 -0.74 -0.42 -0.81 -0.42 0.70Cu -0.13 -0.74 1.00 0.69 0.79 0.79 -0.63Pb -0.41 -0.42 0.69 1.00 0.70 0.80 -0.37Zn -0.09 -0.81 0.79 0.70 1.00 0.75 -0.48Fe 0.80 -0.42 0.79 0.80 0.75 1.00 -0.59Mn 0.86 0.70 -0.63 -0.37 -0.48 -0.59 1.00

Bold values significant to P <0.05.

Linear correlation coefficient between the concentrations of heavy metals in each horizon was calculated in order to assess the presence of any common behaviour, as listed in Table 12. The concentrations of iron, lead, zinc and copper are significantly correlated, according to their calcophilous behaviour, since they all tend to form compounds with sulphur. In this case, these elements are all present in the mineralized ore body of Valle Imperina in the form of pyrite (FeS2), chalcopyrite (CuFeS2), sphalerite (ZnS) and galena (PbS). Most of iron in soils of the study area derives from the alteration of pyrite and chalcopyrite, and this fact explains its low correlation with manganese, even if usually Mn tends to accompany Fe, due to their similar geochemical behaviour. The same combination of elements present in the mineralization of Imperina Valley is found in mine soils. These elements in the soil tend to have similar behaviour, and have a limited mobility, especially in oxidizing conditions. Accordingly, no element of the mineralization has been removed in a preferential way, and this is due to the fact that the agents of pedogenesis have acted for a short time (some decades) in the areas affected by processing the ore material, and thus the chemical characteristics of soil are still tied to that of the substrate. Recent studies by Bini et al. (2011), however, show that several

66


factors may contribute to trace elements concentration in the A horizon. Yet, surface horizons are generally enriched in organic matter, which could have a role in adsorbing trace elements. On the other hand, it is possible that heavy metals be accumulated at surface since the past century, when mining activity was operating in the area. Exploitation, grinding and roasting of minerals could have generated solid particulate added to soil. A third possibility is that metals could have a partial natural (geogenic) origin, and a partial anthropic origin, and the observed stratification could be a result of these two forms of diffused contamination. Moreover, it is likely that higher concentration of trace elements could be related to the migration of species released by the mineralised area, via riverine transport, in the extreme parts of the valley.

67

Chapter 5

DISCUSSION

The present review of some of the many sites of former mining activity in Italy shows that the abandoned waste dumps contain significant amounts of heavy metals that are potentially harmful to the environment.

As reported by Benvenuti et al. (1999) and Mascaro et al. (2001a, b), the main factors controlling the release of toxic elements are:

The original content and composition of the metal-bearing rocks and the buffering phases in the waste. In particular, where pyrite is abundant, as at Boccheggiano and Fenice Capanne, the acidity produced by its oxidation rapidly uses up the buffering phases such as carbonates and aluminosilicates, whereas where galena and sphalerite are dominant, as at Bottino, alteration occurs under near-neutral conditions.

The morphology and particle size of the mine waste. Waste dumps are typically unsaturated with respect to water, and have relatively high hydraulic conductivity that favours oxidation. and influences the alteration processes.

Unlikely, in the tailings impoundments infiltration of oxygenated waters is limited by the low effective porosity of the fine sandy or silty-clayey material, and by the presence of a shallow water table in the wet season. For example, at Fenice Capanne, flotation tailings deposited in unconfined dumps show more advanced alteration than similar material in the impoundments. The particle size of material in the waste dumps influences the alteration processes: all other factors being equal (mineralogy, age, etc.),

coarser grained tailings (pebbles and gravel) show less advanced alteration than sandy-silty materials.

The age of waste pile. This factor seems to be of less importance than the preceding two, but it may become significant in the long term (decades or centuries), as it appears from soils developed from chronologically different waste dumps at the Temperino mine (Bini and Gaballo, 2006). At Bottino, for example, the alteration processes in Renaissance-age waste dumps are significantly more advanced than in dumps from the past century.

The results available for the Boccheggiano area (Benvenuti et al., 1999) suggest that pollutant transport over long distances occurs as suspended particulate matter in streams, and that gravitational runoff, aeolian transport, and transport in solution are limited to the immediate proximity of the sources. There is evidence that a portion of the metals is fixed in relatively stable solid phases, either in specific primary or secondary mineral species, or as minor elements in solid solution; a portion is also adsorbed onto clay minerals and iron-oxyhydroxides. However, metal fixation in the easy soluble or poorly crystalline minerals is ephemeral, and the metals may be released because of geochemical changes related to meteoric events, bacterial activity, or photochemically induced redox reactions (McKnight et al, 1988). Unlikely, in the Imperina Valley district aeolian transport and suspended particulate matter in streams seem to have an important function in disseminating metal pollution in areas conterminous to the mine district (Bini et al., 2011).

Soils influenced by mine spoils present little morphological evidence of profile evolution. Soil parent material may be transformed into immature soils in a relatively short time, as it was found by Néel et al. (2003) in 35-year-old sulphide mine tailings. However, data presented indicate relevant differences in soil development at various sites, where parent material is the main soil forming factor. Evidence of the spatial variability of soils is given by the uneven distribution of vegetation coverage. This is related to the spreading of mine tailings rich in phytotoxic heavy metals (Cu, Pb, Zn) on the land surface: the shorter is the time passed, the slower is natural land revegetation. The processes of soil formation have been driven by the nature and properties of the spoil parent material, such as mineralogy and chemistry, grain size, porosity, etc. Specific processes consist of weathering, oxidation, leaching, and humus

Discussion

addition. As a consequence, a new kind of soil started to form, first as a discontinuous soil cover with A-C profile, and subsequently with a well-developed surface A horizon enriched in organic matter, characteristic of Entisols in Soil Taxonomy (Soil Survey Staff, 1999). Similar Entisols with up to 10-cm thick A-C sequence are known to develop in less than 100 years from natural unconsolidated deposits (Forth and Turk, 1972; Bini et al., 2004), and from mine spoils (Néel et al., 2003). At the sites investigated, the time of exposure to weathering varies from 2700 years to the present, and most soils investigated show little development. Soil thickness increases with time gradually, while plant roots explore thicker layers with increasing density; a thick (up to 100 cm) A-B-C sequence with a cambic horizon may form, giving origin to Inceptisols. However, according to the new proposed classification (ICOMANTH, 2005), all these soils originated from anthropogenic material, and therefore may be called Anthrosols (Technosols in the new ICOMANTH circular letter, 2007).

Combining these results with historical and archaeological data, soil characteristics, and literature on soil genesis in Mediterranean areas (Mirabella et al., 1992; Bech et al., 1997; Bini et al., 2006), the following tentative scheme of soil evolution (Table 13), corresponding to a provisional chronosequence (although based on relative age), may be suggested for Anthrosols containing mine spoil as a parent material in the profile, and for the corresponding “normal” soils:

Table 13. Tentative scheme of soil evolution, and correlation between Anthrosols and “normal soils” in the study area

Estimated age

Anthrosols “Normal” soils

<100y (LITHIC) SPOLIC XERORTHENT

TYPIC XERORTHENT

<1000y SPOLIC DYSTROXEREPTSPOLIC HAPLOXEREPT

TYPIC HAPLOXEREPT

> 1000 y SPOLIC RHODOXERALF TYPIC RHODOXERALFSPOLIC XERORTHENT over TYPIC RHODOXERALF

Anthrosol taxonomy from ICOMANTH (2005).

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Claudio Bini

As stated by Bini and Zilocchi (2004), the occurrence of soils at different stages of development (chronosequences) enables the establishment of chronofunctions related to a given time interval (2700 years in the present study). The soil-time function (or chronofunction) is defined by the following equation (Jenny, 1941, 1980):

s=∫t0

t n

(cl , o ,r , p )

in which s=soil, t=time, cl=climate, o=organisms, r=relief, p=parent material.

Several soil properties have been utilized, and numerous examples of chronofunctions have been reported in the literature (e.g. Harden, 1982; Bockheim, 1980, 1990; Schaetzl et al., 1994; Rabenhorst, 1997). In a recent study, in order to evaluate the effect of time on soil formation from mine dumps, Bini and Gaballo (2006) tried to find out different soil properties (SOC, colour, A horizon thickness, total thickness, pH, particle size) to relate to soil age, and to determine a possible chronofunction for such soils. Among the investigated soil properties, the most effective in defining an actual chronofunction proved the A horizon thickness. We have measured the thickness of the A horizon of the examined profiles, and we have plotted it against the relative (archaeological) age of mine spoil. The A horizon thickness increases linearly with increasing soil age, as expressed by the equation:

A horizon thickness (cm) = 0.011(soil age in yrs).

The correlation coefficient (R2) is 0.9766, indicating that nearly all variation in soil thickness is a function of soil age. This suggests that organic matter addition and humification are key processes in the first steps of soil development, and, therefore, vegetation has a key role as initial soil forming factor. Moreover, although it is generally difficult to quantify the effect of time on pedogenesis (Bockheim, 1980; Yaalon, 1997), a single soil property (e.g., A horizon thickness, organic carbon content) can be effective to build up a simple chronofunction, at least over short time intervals.

72

Discussion

In conclusion, the comparative study of Anthrosols and “normal” soils (unaffected by mine spoil) allowed reconstruction of the natural and anthropogenic stages of soil development, as well as the identification of the main governing factors.

Mine tailings may be considered colluvial deposits whose discharge at surface caused different trends in pedogenesis and noteworthy spatial variability. Current soil development is governed by the composition and properties of mine spoil. The age of spoil, combined with inherited effects of the spoil are responsible for different pedogenetic stages; tailings are likely to generate different types of soils (Anthrosols) over short distances (<0.5km) from mine waste.

The original, “normal,” soil (paleosol) is truncated and/or buried by potentially phytotoxic spoil. The anthropic intervention in the mining area and in the conterminous land impacted heavily the landscape and the natural environment. About 1/6 of the territory examined has been impacted by mine spoil, with effects such as changed surface topography, immature infertile soils, and poorly structured, herbaceous and shrubby vegetation coverage replacing the original Mediterranean maquis.

According to this, and compared with results obtained in other contexts (Néel et al., 2003; Bini and Zilocchi, 2004), the recorded trend of soil evolution contributes to a better knowledge of areas affected by similar waste material, and may be utilized in remediation of abandoned mine areas.

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Chapter 6

SUMMARY AND CONCLUSIONS

Existing literature shows that there are not significant differences among the various mine sites investigated, irrespective of the nature of ore deposits. The abandoned mine waste contain significant amounts of polluting elements potentially dangerous to the environment. The main differences concern the waste morphology (dumps, tailings, roastings, soil), the particle-size, the hydrological regime, the pH conditions, which determine the fate of metals in the environment.

In particular, literature summarised in this review indicates that:

Sulphides that occur in flotation basins show comparatively low degrees of alteration in comparison to exposed mine dumps: scarce effective porosity, shallow water table and superficial capillary fringe have hindered sulphide weathering; near neutral pH and secondary mineralogical phases indicate that until now reactions producing and consuming acidity balance each other;

Inversely, the hydraulic unsaturated conditions and the fine size of flotation tailings have enhanced the weathering of sulphides;

The roasting dumps are characterised by high contents of metals and low pH values. However, they appear to be a less significant threat to the environment, at least in the short term, because the metalliferous phases are partially enclosed within relatively coarse quartz grains, and adsorbed onto alumosilicates, thus slowing the weathering processes;

Stream water shows acid to neutral pH values, and maintain high metal contents as far as 1 km downstream of mine wastes; more

distal waters are neutral as a consequence of dilution with unpolluted water and buffering reactions with carbonate bedrock;

The transport of polluted stream sediments causes contamination as far as 5 km downstream of the mine wastes.

Aeolian transport and suspended particulate matter in streams seem to significantly contribute in disseminating metal pollution over large areas, even at distance from the metal source.

The results presented shed light on the environmental effects of mine wastes, give a thorough understanding of the polluting potential of mine waste, and can be an useful basis for planning possible remediation projects. Yet, the variation of the above cited conditions could cause the establishment of more acidic and leaching conditions and, therefore, an increase in metal mobilization, and major environmental hazard.

Based on available data, a possible remediation plan could include the preservation of the existing conditions, enlarging the impoundment surface by building of settling ponds for drainage waters storage. Afterwards, the polluted sediments could be stored into the flotation basins, and mixed with buffering compounds such as limestone for neutralisation, and prevent acidic drainage waters production. Finally, revegetation of the whole mined area with metal-tolerant plants could take place, and the contaminated land be restored.

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Chapter 1 - Home page | Archivio Istituzionale della Ricerca · Web viewAn absorption sequence Zn>Co>Cu>Ni>Fe>Cr, consistent with leaching tests, was found by Dinelli and Lombini (1996)