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«Dissertation zur Erlangung des Grades “Doktor der Naturwissenschaften” Im Promotionsfach Geowissenschaften Am Fachbereich Chemie, Pharmazie und ...»

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Soils can be viewed as an inhomogeneous medium with the humus layer serves as a natural biogeochemical barrier that suppresses the percolation of the element with the seepage water, an as a result strongly accumulates the element. High As concentrations in natural soils are usually associated with sulfide minerals (chalcophilic) and their weathering products. The major As-bearing minerals are mixed sulfides of FeAsS, NiAsS, CoAsS as well as other twovalent metals (e.g. FeAsS2, AsS, NiAs and CoAsS). All these minerals are generally believed to be formed under high temperature especially like arsenopyrite can be derived from hydrothermal solution at 100oC or more of conditions in the earth’s crust. The rules which governing the chemistry and mobility of arsenic in soils are depending on several factors such as soil solution chemistry, pH, redox conditions, soil solid composition, As-bearing phases, adsorption and desorption, biological transformations. The soil constituents related to As mobility are usually oxides of Fe, Al and Mn, clay minerals, and organic matter. Arsenic in soils may distribute among various soil components in different physicochemical forms which are associated with various soil constituents. The prime concern is the chemical associations of As with various soils solid phases instead of its total concentration which affects its mobility, bioavailability and toxicity to the biosphere. The contents of Si, Al, and Fe reflect the intensity of the transformation of primary minerals and the formation of clay ones. Compounds of biophilic elements characterize the fertility and supply of soils with available nutrients. The distribution of pollutants between soil components determines the environmental consequences of soil pollution.

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(1.1.3) Arsenic in minerals Arsenic occurs in most minerals such as elemental As, arsenides, sulphides, oxides, arsenates and arsenite. A list of some of the most common As minerals can be shown in following table (2), (Smedley 2002).

Table (2) Table shows major As minerals in nature

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The most abundant As ore mineral is arsenopyrite (FeAsS) with the chemistry of As follows closely with sulphur. Arsenic can also be found in common rock-forming minerals with greatest concentrations of the element in sulphide minerals in which pyrite is the most abundant.

Arsenic is usually involved with crystal structure of many sulphide minerals as a substitute for S.

High As concentrations can also be found in many oxide minerals and hydrous metal oxides, either as part of the mineral structure or as sorbed species. It is also noticed that arsenic presents in phosphate, silicate and carbonate minerals although their abundances and concentrations are far less compared with oxide minerals.

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(1.1.4) Toxicity and mobility of arsenic Most of the arsenic compounds are toxic such as arsine gas, arsenic trioxide, sodium arsenite and sodium arsenate. Both of the As3+ and As5+ area capable of inhibiting the energylinked functions of the mitochondria and also causing inhibition of DNA damage repair. Chronic arsenic poisoning can arise numerous symptoms such as anorexia, hyperkeratosis, cardiovascular diseases and etc. Acute arsenic poisoning may lead to muscle cramps, abdominal pain, renal failure and etc. Human body to certain amount can detoxify the inorganic As 3+ and As5+ compounds by methylation.

Eh-pH diagram for aqueous As species in the system As-O2-H2O at 25 oC and 1 Fig. (1) bar total pressure (Section 8.1 Ref. 67) The above figure (1) shows redox potential (Eh) and pH as key factors in controlling As speciation. For example H2AsO4 - is a dominant species at low pH (6.9) under oxidizing conditions. On the other hand, HAsO42- becomes dominant in higher pH. Uncharged arsenite species H3 AsO30 predominates in reducing conditions with pH less than 9.2.

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Fig. (2) Arsenite and arsenate speciation as a fuction of pH (ionic strength of about 0.01 M). Redox conditions have been chosen such that the indicated oxidation state dominates the speciation in both cases. (Section 8.1 Ref. 67) The above figure (2) shows the distribution of As species as a function of pH. It is however important to notice that degree of protonation is an important factor in determining the speciation.

Under high concentrations of reduced sulphur can cause significant present of As-sulphide species and on the contrary precipitation of orpiment and realgar favors in reducing acidic conditions.


Species of soil compounds and their contents, the pH value and the redox potential all play an important role for the mobility of arsenic species in soils. Arsenic compounds can adsorb to oxides and hydroxides of Fe (III), Al (III), Mn (III/IV) and clay minerals. High redox potential and acidic condition can keep arsenic out of mobilization. Arsenic can be mobilized when Fe 3+ and Mn3+/4+ are reduced. The adsorption of arsenic by soils depends on the content of amorphous iron oxides and the five most important Fe (III) oxides/hydroxides are Fe(III) hydroxide (Fe(OH)3), goethite (α-FeOOH), akaganeite (β-FeOOH), lepidocrocite (γ-FeOOH), and haematite (Fe2O3). The fixation of arsenic is highly influenced by the specific surface and crystallinity of the Fe oxides. As5+ can form inner-spherical bi-dentate surface complexes and monodentate surface complexes can also be formed when adsorbing to Fe(OH)3. Anionic arsenic compounds can be fixed by hydrated Fe3+ oxides in aquatic environment. In addition, arsenic can also co-precipitate with Fe(OH)3. Moreover, sorption of arsenic in soils can be depended on the presence of amorphous hydrated aluminum oxides, Mn oxides present in soils, existence of calcium ions (e.g. change of the surface charge characteristics of the soils) on As5+ adsorption.

Furthermore, phosphates can suppress As3+ and As5+ sorption by soils especially when soils contain low amounts of iron oxides (Bissen, 2003).

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(1.2.1) Introduction of copper Copper is a metal existing in different oxidation state (0, +1 and +2) with +2 the most common one. Copper contains the two stable isotopes Cu and Cu. It is an essential trace element for all the living organisms including humans, plants, animals and micro-organisms.

Copper can be found naturally in most soils, fruits and vegetables. Human requires regular intake of copper with the element playing key roles in the production of blood haemoglobin. Copper in plant is used for seed production, disease resistance and regulation of water. In natural environment copper is relatively abundant in the earth crust with most of the copper occurring in unavailable mineral form. The worldwide emission of copper from natural sources were estimated and showed at the follows (table 3) (modified Pacyna 1986).

Table (3) Table shows worldwide emission of copper from natural sources

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There are numerous human activities which involve usage of copper such as textiles, electrical conductors like copper wire and electromagnets, coins and cooking utensils. Copper can be alloyed with nickel (e.g. cupronickel) and used as corrosive resistant materials in shipbuilding. Copper compounds are active ingredients of pesticides and fungicides (e.g.

Bordeaux mixture: CuSO4.5H2O + Ca(OH)2). Copper sulphate is the most common source of fertilizers for copper-deficient soils where the soils are used for growth of vegetables. CuO and mixtures of CuSO4 and Cu(OH)2 are also used as copper micronutrients in agriculture. Copper based compounds such as chromate copper arsenate (CCA) were widely used as wood preservatives. Copper according to EMRC (1992) was the fifth most-valued mineral commodity produced in Canada and accounted for 5.8% of the total value of Canadian mineral production.

Copper occurs in a wide range of mineral deposit as primary and secondary minerals.

Most copper occurs in the form of sulphide minerals chalcopyrite (CuFeS2), chalcocite (Cu2S), bornite (Cu5FeS4) and tetrahedrite ((CuFe)12Sb4S13). During chemical weathering the primary copper sulphide minerals, secondary minerals maybe formed cuprite (Cu 2O), malachite (Cu2(CO3)(OH)2), azurite (Cu3(CO3)2(OH)2, brochantite (Cu4SO4(OH)6), antlerite (Cu3SO4(OH)4.

The following table (4) showed the range of copper concentration (ppm) in igneous and sedimentary rocks.

Table (4) Table shows copper concentrations in igneous and sedimentary rocks (modified Cannon et al. 1978)

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The association form of copper can be affected by various environmental conditions such as pH, temperature, redox potential, organic matter decomposition, leaching and ion exchange processes and microbial activity. The availability of copper for biological uptake and transport in the environment is controlled by processes at solid-water interfaces and copper speciation in solution. Copper sulphide minerals can be oxidized and the resultant Cu 2+ ion can be released in complexed hydrated form such as Cu(H2O)62+. In aqueous phases copper is predominantly in the more stable Cu2+ state and when in contacts with water copper can form [Cu(H2O)6]2+ ion and further complexes with inorganic and organic ligands. The final copper complexes can be absorbed onto clays, sediments and organic particulates. pH, copper concentrations, competing cations and the adsorbent’s properties are the factors influencing copper adsorption. Desorption of copper can also occur with the present of Ca2+ and Mg2+ competing cations. Most of the copper found in natural water is partitioned in sediments. The physiochemical processes determine the transport and distribution of copper and its complexes in soil are adsorption, aqueous-phase solubility, leaching and lateral movement (aeolian and fluvial). Minimal copper can be lost through volatilization and the copper complexes can be broken down to copper ions by physical and biological processes. All these processes control dissolved copper concentrations, copper mobility and bioavailability in the environment. The copper ions in solution can be complexed, precipitated and adsorbed. Copper usually exists in CuO and Cu2O in high temperature combustion to the atmosphere. Copper in atmosphere can be removed by bulk deposition and wet/dry deposition mechanisms.


(1.2.2) Copper concentrations in unpolluted soils The amount of copper in soil is depending on the parental rock, distance from natural ore bodies and manmade air emission sources. The copper content in soils is commonly range from 2 to 100 mg/kg (Allaway 1968) with an average value of 30 mg/kg. Various anthropogenic and natural sources such as copper mining, metal smelters, waste incineration, agricultural and industrial application can contribute to the presence of copper in atmosphere and soil. Copper concentrations in soil can vary between soil type, soil amendments, distance from anthropogenic sources, distance from natural ore bodies and composition of bedrock and parent material.

According to the ministry of the environment Ontario, Canada the background concentrations of copper in Ontario soils are on average less than 25 mg/kg with exceptional circumstances at 85 mg/kg. The average copper concentration in Canadian soil is estimated to be 20 mg/kg with a range between 2 and 100 mg/kg (British Columbia Ministry of the Environment, Lands and Parks 1992). Reave et al. 2006 reported the total copper contents of 4179 samples from 946 Scottish soil profiles showed the derived mean of copper concentration from all the samples is 10 mg/kg and the normal copper concentration range is between 0.93–110 mg/kg. The copper contents of Tibetan soils were recorded by Xiaoping et al.2002. The collected 205 soil samples from the remote area of Tibet Plateau in 5 soil classes (0–20 cm) were analyzed with the average content of Cu was 19.6 mg/kg (CV=49. 28%). Tibet is one of the few places in this world that are least influenced by human activities and an ideal location of investigating natural background copper concentration and can possibly serve as a reference source for the world. McKeague et al.

(1979) reported the highest mean copper concentrations were in Cordilleran region (46 mg/kg) whereas the Canadian Shield contained only a mere 12 mg/kg. The mean copper concentration from all regions in Canada was 22 mg/kg (McKeague and Wolynetz 1980). Mean copper concentrations in rural parkland and old urban parkland surface soils samples (0-5 cm) were found to be 16 and 26 mg/kg respectively by the Ontario Ministry of the Environment and Energy. The city of Trondheim in Norway conducted a soil geochemistry survey in 1994 and 2004 by analyzing 321 surface soil samples (Andersson et al. 2010), the average copper concentrations were found to be 42 mg/kg and 39 mg/kg respectively. In the Peace River, Alberta the surface and subsurface soils mean total copper concentration was found to be 22.1 mg/kg with samples ranging from gleysols, Luvisolic and Solonetzic soils to Podzolic soils (Soon and Abboud, 1990). Lai et al. (2010) conducted a field investigation to evaluate copper


contents in vineyard soils at central Taiwan and the copper concentration was range from 9.1 to 100 mg/kg. Soils samples collected from A and C horizons in southern and western Manitoba were in average of 25 mg/kg (CCME 1997). In southwestern Ontario the soil samples collected from agricultural watersheds, the soils from various horizons (Ap, B and C) showed mean total copper concentrations no more than 27.4 mg/kg (Whitby et al. 1978).

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(1.2.3) Copper concentrations in polluted soils A strong correlation was found between base metal levels in peat and distance from the base metal mining and smelter complex in Flin Flon, Canada (Zoltai 1988). Surface soils in the vicinity of copper smelters were found to be heavily contaminated by atmosphere fallout.

Average total copper concentrations were often well above 1000 mg/kg (Hazlett et al. 1983;

Hutchinson and Whitby 1974; Kuo et al. 1983). At Coniston, near Sudbury, Ontario, maximum concentrations of copper reached 9700 mg/kg within 0.41 km of the smelter site (Hazlett et al.

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