Download Consequences of Mining Operations on Environmental

ENVIRONMENTAL IMPACT OF
ABANDONED MINE WASTE:
A REVIEW
ENVIRONMENTAL IMPACT OF
ABANDONED MINE WASTE:
A REVIEW
CLAUDIO BINI
AUTHOR
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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
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Claudio Bini
Museums; Archaeological Parks; protected natural areas, didacticrecreational 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
3
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 riskbased 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 15th
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,
Tunisia) began in Etruscan times, was expanded by Romans and
dominated the world mercury production for long time (Bargagli, 1995).
4
Claudio Bini
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 (copperbronze 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
antimony
arsenic
cadmium
chromium
copper
lead
mercury
nickel
zinc
Data is in metric tonnes x103 / year.
Modified after McGrath, 1995.
Mine production 1990
55
45
19
6800
8110
3100
6.8
778
6040
Introduction
5
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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Claudio Bini
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 (HeliosRybicka, 1996; Cidu et al., 2009). Smelting of ore deposits results in the
release of metals to the atmosphere (Mihalik et al., 2011); when metals
have been released through the atmosphere, they end up as diffuse
pollutants in soils and sediments. (Nriagu, 1990; Salomons, 1995).
Introduction
7
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
gradients); vegetation cover may have great influence on attenuating, or
enhancing, erosion phenomena, when land cover is scarce or lacking, as it
8
Claudio Bini
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
frequently recorded in workers employed in mining industry and even in
hat factories in Tuscany (Dall’Aglio et al., 1966). At present, diseases and
Introduction
9
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 cardiovascular 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 20th 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 Cr VI 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
as a tolerable level in drinkable waters, therefore, has been recently
proposed by the World Health Organization (2004).
10
Claudio Bini
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).
Introduction
11
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 protohistorical 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 (16th-19th
century), ore exploitation has been carried out by both local population and
12
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 metalcontaminated soils, and restoration plans have been addressed to these
sites, with application of low cost and environmental friendly
phytoremediation technologies (Bini, 2009; 2010).
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
14
Claudio Bini
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
15
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++4H2OMe2++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
16
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
Processes Occurring at the Mine Sites
17
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.
18
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;
Processes Occurring at the Mine Sites
19
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
20
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 (SO42-, Fe, Zn, Pb, As and REE) were found also
Processes Occurring at the Mine Sites
21
in surface waters up to 1500m downstream from a mine site in Cuba
(Romero et al., 2010).
The mineralogy of overbank sediments is mainly composed of quartz,
feldspar, calcite and phyllites (i.e. clay minerals) as principal phases; the
largest quantities of pyrite, chalcopyrite, galena, sphalerite, Fe oxides
occur generally close to the waste and decrease downstream (Mascaro et
al., 2001a). The ochreous muds frequently occurring at these sites consist
of iron oxhydroxides (ferrihydrite) and quartz mechanically transported as
suspended matter during flooding episodes; it is likely these muds to be
responsible for the high metal content found in bottom sediments and in
soils along the river overbank. As a general rule, bulk chemistry of
sediments seems to be influenced by proximity to the mine waste; indeed,
as far as 5 km downstream, sediments maintain relatively high contents of
metals, in particular Zn.
Figure 5. Overbank sediments along the Imperina creek (Belluno, Northern Italy).
(Photo Bini).
22
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).
Processes Occurring at the Mine Sites
23
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; MorenoJimenez 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).
24
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 wellexpressed 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;
Processes Occurring at the Mine Sites
25
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
26
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.
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
28
Claudio Bini
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
Depth
Horizon
cm
A1
A2
2C
0-10
10-30
30-70
Particle size pH Total
Organic Organic CEC
(USDA)
Carbonates Carbon Matter
%
silt clay sand
g/kg
g/kg
g/kg
cmol(+)/
kg
25 14.8 60.2 7.8 0
16
28
20
26 27.4 46.6 7.6 0
14
24
22.5
32.7 37.5 29.8 7.5 0
5
9
14.5
Consequences of Mining Operations ...
29
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).
30
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.
50
51.6
58.1
6.4
6.3
6.2
g/kg
0
0
0
27
21
8
46
36
13
cmol
(+)/
kg
25.5
15.4
13.0
cmol
(+)/
kg
15.6
17.4
9.3
Base
Saturation
g/kg
Exch. Acid.
15.3
15.1
11.1
g/kg
CEC
sand
Organic
Matter
0-47 34.7
47-70 33.3
70-90 30.8
clay
Organic
Carbon
silt
pH
Particle size
(USDA)
%
cm
Total
Carbonates
A1
A2
Bw
Depth
Soil
Horizon
Table 3. Selected properties of the Spolic Dystroxerept illustrated in
Figure 9
%
62
47
58
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).
31
32
Claudio Bini
silt
clay
sand
0-3
3-15
15-40
40-110
110-120
23,9
29,8
35,7
40,7
39,9
9,5
13,7
9,9
40,3
30
66,6
56,5
54,4
19
30,7
7,6
7,7
7,9
8,0
8,0
g/kg
g/kg
g/kg
0
0
0
4,1
58
23
19
6
7
11
40
33
11
12
19
CEC
Organic
Matter
pH
cm
O. C.
A1
A2
E
Bt1
Bt2
depth
hori-zon
Particle size USDA
Carbonate
Table 4. Selected properties of the Spolic Haploxeralf
illustrated in Figure 10
cmol(+)/
Kg
29,5
37
33
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).
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
34
Claudio Bini
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.
6
4
1 32
5
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.
Case Studies in Italy
35
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
36
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
Mine
22
0.2
730
336
919
waste
Mine soil
18
0.09
230
146
634
(1995)
Proximal
12
0.07
120
164
451
soil
(1995)
Distal soil 8
0.07
62
131
401
(1995)
Source: Corsini et al., 1980; Bini et al., 1995; Servida et al., 2009).
Fe, Mn are expressed as %; Cu, Pb, Zn as mgkg-1.
pH
texture
n.d.
gravel
5.3
Gravel,
sand
Sandy
loam
6.3
6.6
Loamy
sand
Case Studies in Italy
37
4.1.2. The Metalliferous Hills District (Massa Marittima)
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
38
Claudio Bini
magmatism, which were responsible also for the alteration and replacement
of host rocks, with the formation of skarn silicates (mainly pyroxenes,
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 (6th7th 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 19th
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.
Case Studies in Italy
39
The mineralogy of mine waste has been recently studied in detail by
Mascaro et al. (2001a). Quartz and hematite, partly derived from roasting
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 illitemontmorillonite 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).
40
Claudio Bini
Table 6. Average concentrations of heavy metals, pH and grain size at
Fenice Capanne
Sample Fe %
As
Bi
(age)
Roastings 17
927 320
(ca.1890)
Tailings 3.6
515 22
(ca.1950)
Tailings1 7.7
230 84
(ca.1960)
Tailings2 5.1
116 30
(1970)
Proximal 5.8
123 28
soil
(1999)
Distal
6.3
119 27
soil
(1999)
Adapted from Mascaro et al., 2001a.
Cu
Mn
Pb
Zn
pH
Texture
3,450 418
825
2,480 3.5
853
2,050 2,490 4.2
Gravel,
pebble
Sand,
silt
Sand,
silt
Silt,
clay
Coarse
loamy
424
3,400 4,320 1,310 5,470 7.1
508
3,520 214
1,560 7.4
595
3,700 353
1,720 6.4
445
2,900 413
522
6.7
Coarse
loamy
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
Case Studies in Italy
41
(Cu(Pb-Zn-Ag)) ores, but then, from 1906 to 1994, some tens of millions
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.
42
Claudio Bini
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)
Mine
waste
(ca.1889)
Mine
waste
(ca.1910)
Mine
waste
(ca.1950)
Tailings
(1970)
Fe
12.8
(1.619.3)
13.6
(4.724.5)
10.4
(9.411.3)
9.4
(5.215.2)
Proximal 10.1
soil
(1.6(1995)
15.9)
Distal soil 4.2
(1995)
(1.65.8)
As
Bi
188
(13264)
241
(88429)
737
(551000)
233
(74887)
640
(50900)
55
(40350)
153
(37339)
26 (480)
Cu
Pb
589
590
(529.642) (2771110)
754 (36- 3160
1790)
(3428000)
29
196 (29- 304
(14- 468)
(10640)
956)
13 (3- 203 (71- 2145
44)
377)
(4494920)
17 (2- 81 (35- 402
29)
800)
(110956)
10 (1- 59 (29- 81(5014)
206)
890)
Zn
pH
Grain size
292
3.1
(234327)
758 (77- 4.2
1970)
Gravel,
pebble
392 (40- 4.3
1300)
Pebble,
sand
4980
5.5
(1549900)
312 (97- 4.8
446)
Silt, clay
176 (40- 6.7
331)
Coarse
loamy
Gravel,
pebble
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
Case Studies in Italy
43
(range: Zn = 150-9900 ppm; Pb = 450-4920); Cu and Bi in the roasting
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.
44
Claudio Bini
Soils
The 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
Case Studies in Italy
45
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-PbZn 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
46
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, FeCu 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
Case Studies in Italy
47
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 Dumps
The 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.
48
Claudio Bini
Soils
The 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
Case Studies in Italy
49
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 5.17.9), 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 Geochemistry
Waste 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
50
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.
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
Na2O MgO
Horizon
Spolic
Xerorthent
Waste soil
Spolic
Dystroxerept
1.10
0.90
0.43
0.49
0.18
0.62
0.19
0.21
0.18
0.18
0.02
0.21
0.22
0.35
0.31
0.05
A1
A2
2C
A1
A2
Bw
Proximal soil 2Bwb
2Bwb’
3BCb
3Cb
Spolic
A
Xerorthent
E
Over
EB
Rhodoxeralf 2Bt1
Distal soil
2Bt2
3C
1.45
1.66
3.36
2.04
1.39
2.99
1.46
3.09
3.73
7.35
4.26
4.55
3.86
2.69
2.45
11.52
Al2O3
SiO2
P 2O 5
K2O
CaO
TiO2
MnO
Fe2O3
Cu
Pb
Zn
12.53
15.23
12.70
6.60
3.87
10.52
13.72
13.75
16.62
22.16
15.15
15.72
16.42
18.47
24.33
22.67
72.64
66.35
41.54
43.57
47.35
37.53
37.17
64.41
65.94
49.14
45.95
60.29
58.84
60.35
52.82
45.02
0.06
0.06
0.40
0.05
0.04
0.07
0.03
0.08
0.04
0.05
0.16
0.06
0.07
0.06
0.05
0.04
2.07
2.39
1.70
1.23
0.69
1.76
0.66
2.54
2.68
3.04
3.05
2.84
2.95
3.34
3.21
3.72
0.63
0.62
6.47
3.32
4.55
1.73
6.22
4.73
0.50
0.69
4.75
0.89
0.90
0.65
0.64
0.89
0.63
0.69
0.73
0.21
0.12
0.41
0.21
0.54
0.56
0.70
0.79
0.73
0.71
0.81
0.77
1.00
0.19
0.24
4.36
3.91
3.10
3.16
3.92
1.30
0.89
2.83
1.50
1.03
1.18
0.95
1.04
0.57
6.66
8.15
14.70
33.81
44.46
23.78
41.33
10.61
6.38
9.61
10.38
9.83
10.04
9.21
10.91
9.98
170
260
1000
9665
14100
3790
14900
7310
3700
9230
420
280
590
390
140
16
78
24
840
9200
4100
11900
3770
1100
340
320
420
460
1500
470
330
65
260
120
2000
15000
5840
19800
6240
2640
2400
4300
920
620
1260
780
510
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).
Mine
waste
(1593)
Mine
waste
(1840)
Mine
waste
(1929)
Mine soil
(1995)
Fe
Mn
Pb
Zn
S
As
Cd
Cu
Sb
pH
texture
9.8 (6.912.5)
0.6 (0.30.9)
0.5
(0.05-1
0.7 (0.50.8)
86 (3174)
0.4
(0.050.7)
0.3 (0.10.6)
1.0 0.24.9)
75 (3150)
30
(2438)
75 (5485)
1.0 (0.42.7)
266
(95576)
470
(100760)
112
(90130)
70
(5080)
442
(272568)
177
(40430)
384
(133983)
188
(70320)
52
(3570)
23 (633)
110
(3170)
82
(18260)
224
(41680)
700
(1101750)
16 (728)
6.8
(6.57.3)
6.8
(3.57.9)
6.2
(4.76.6)
5.7
(5.36.1)
5.1
(4.95.3)
4.6
(4.05.0)
Gravel,
pebble
7.6 (4.110.4)
0.3
(0.30.4)
0.6
(0.013.2)
0.7
(0.12.5)
0.9
(0.31.3)
0.2
(0.10.2
0.2
(0.20.3)
0.6
(0.063.9)
7.8 (6.80.8
9.5)
(0.22.0)
5.7 (7.20.3 (0.30.7
8.7)
0.4)
(0.22.4)
Proximal
5.6 (5.00.1
0.05
soil
6.6)
(0.07(0.02(1995)
0.1)
0.1)
Distal soil 4.9 (3.00.02
0.03
(1995)
5.8)
(0.01(0.020.03)
0.03)
Adapted after Mascaro et al. (2000; 2001b).
1.5 (0.43.1)
0.04
(0.010.2)
0.03
(0.010.03)
45 (5140)
60
(2490)
6 (59)
7 (511)
7 (58)
Gravel,
pebble
Gravel,
pebble
Sandy
skeletal
Sandy
loam
Loamy
sand
Claudio Bini
53
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 quartzcarbonate-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
tourmaline. Carbonates are mainly represented by terms of the siderite-
54
Claudio Bini
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)
b)
c)
d)
submicroscopic admixtures of separate phases;
isomorphogenous substitution in the lattice;
surface adsorption;
any combination of the three possibilities.
Minor minerals are ferrihydrite, pyrochroite, cuprite, malachite, Fesulfates, 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).
The main reactions controlling the generation of acidity, which in turn
Case Studies in Italy
55
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↓ + SO42-, 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
alteration. However, given the low amounts of both sulphides and calcite
56
Claudio Bini
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
produced by the forest ecosystem, and the scarcity of acidity buffers such
Case Studies in Italy
57
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;
Trois et al., 2007; Cidu et al., 2009; Da Pelo et al., 2009; Frau et al., 2009),
58
Claudio Bini
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
Caboi et al. (1999) for another mine area. Another possible removal
59
Case Studies in Italy
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 wasterock 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
Mine
waste
Tailings
Fe
Mn
As
Cd
Cu
Pb
Zn
pH
5.9
1.20
225
0.6
982
68
107
3.3
texture
Gravel,
pebble
1.7 0.75
280 <0.5 500 94
41
6.7
Coarse
Loamy
Distal
0.4 0.33
32
<0.5 35
87
56
5.9
Sandy
soils
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.
60
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;
the repetitive presence of metal ions concentration peaks on
resumption of each irrigation, regardless the extent of dry periods;
Case Studies in Italy

61
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 NorthEastern territories, where forests, for building houses and ships, and
metals, smelted for the coinage and the armaments, were easily available.
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
62
Claudio Bini
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
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
Case Studies in Italy
63
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,
favouring the removal from soil and amplifying the risk of water
contamination (Adriano, 2001).
64
Claudio Bini
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)
Prof. 1 waste
Pb
Ni
Cr
Cu
Zn
Fe
Mn
mg
mg kg-1 mg kg-1 mg kg-1
mg kg-1 %
mg kg-1
kg-1
< DL < DL 3159.49 23605.52 1588.92 52.31 256.07
65
Case Studies in Italy
soil
average
range
Prof. 2 waste
soil average
range
Pr. 3 stream
bed average
range
Prof. 4 waste
soil average
range
Pr. 5 stream
bed average
range
Prof. 6 waste
soil average
range
Prof. 7 distal
soil average
range
Prof. 8 distal
soil average
range
-
-
< DL 41.07
2198.19- 20814.97- 1162.91- 50.87- 169.814063.93 28154.16 1786.37 53.55 440.46
2494.35 7057.91 980.02 17.90 506.38
-
20.49- 1936.30- 2497.47- 799.73- 9.66- 279.9959.76 3367.14 14634.89 1188.32 31.86 669.44
54.77 95.31 1122.31 372.98
734.45 6.85 1075.96
49.12- 88.85- 526.44- 227.7760.42 101.77 1718.18 518.18
< DL 39.54 3093.55 5815.92
471.99- 6.18- 985.68996.90 7.52 1166.24
1192.20 40.30 84.09
-
24.26- 1768.54- 1435.79- 271.73- 19.72- 31.8348.28 4419.89 14619.29 2423.03 58.22 174.59
48.64 95.75 512.02 512.02
476.66 5.60 1199.23
46.62- 98.71- 524.29- 205.4650.65 92.80 499.75 293.75
< DL < DL 1639.04 8256.22
-
430.85- 5.91- 1139.51522.47 5.29 1258.96
1338.37 43.23 215.32
-
502.51- 397.04408.52- 4.46- 114.772333.92 12026.83 2566.27 56.82 522.42
27.72 163.63 53.98
64.08
71.23
3.68 811.13
16.91- 147.10- 30.7957.68 177.17 97.91
13.66 31.37
52.1472.43
1224.44 570.79
12.18- 26.85- 282.76- 342.7416.24 34.17 1774.73 762.99
100
51
21
Italian average 46
(1992)
200
20
10
International 40
average (1992)
120
150
120
100
Residential
Limits
(DL 152/2006)
<DL = less than the detection limit.
103.1440.80
3.14- 582.824.02 1024.67
816.79
1.48
181.72
576.80- 1.28- 151.911039.90 1.69 202.07
89
3.70 900
50
-
850
150
-
-
66
Claudio Bini
Table 12. Linear correlation coefficient calculated on the
concentrations of metals in soils in all horizons
Ni
Cr
Cu
Ni
1.00
0.31
-0.13
Cr
0.31
1.00
-0.74
Cu
-0.13
1.00
-0.74
Pb
-0.41
-0.42
0.69
Zn
-0.09
-0.81
0.79
Fe
-0.42
0.80
0.79
Mn
0.86
0.70
-0.63
Bold values significant to P <0.05.
Pb
-0.41
-0.42
0.69
1.00
0.70
0.80
-0.37
Zn
-0.09
-0.81
0.79
0.70
1.00
0.75
-0.48
Fe
0.80
-0.42
0.79
0.80
0.75
1.00
-0.59
Mn
0.86
0.70
-0.63
-0.37
-0.48
-0.59
1.00
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
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
Case Studies in Italy
67
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.
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 nearneutral 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.),
70
Claudio Bini

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
71
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 welldeveloped 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
<100y
Anthrosols
“Normal” soils
(LITHIC) SPOLIC
TYPIC XERORTHENT
XERORTHENT
<1000y
SPOLIC DYSTROXEREPT
TYPIC HAPLOXEREPT
SPOLIC HAPLOXEREPT
> 1000 y
SPOLIC RHODOXERALF
TYPIC RHODOXERALF
SPOLIC XERORTHENT over TYPIC RHODOXERALF
Anthrosol taxonomy from ICOMANTH (2005).
72
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
tn
 cl , o, r , p 
t0
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.
Discussion
73
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.
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;
76
Claudio Bini



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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