Published on International Journal of Agriculture & Agribusiness
Publication Date: June 26, 2019
Mango, L. & Kugedera, A. T
Deparrtment of Agriculture Management, Zimbabwe Open University, Bindura
Deparrtment of Agriculture Management, Zimbabwe Open University, Masvingo
Disposal of mine tailings cause degradation to the ecosystem and a lot needs to be done to improve mining areas. The use of agroforestry species was seen as an option to improve these areas. The use of agroforestry species such as Acacia species, Leucaena and Tephrosia has the mandate to improve mine dumps into useful areas for agriculture and other purposes. These species improves soil fertility by absorbing toxic chemical and metals deposited in the soil from mines. There is great use of improving these areas if management is put in place through responsible authorities. These areas may also be turned into an exciting area where people can come and rest.
Keyword: Mine tailings, ecosystem, agroforestry & Acacia species.
Contaminated land is a problem worldwide, often resulting from industrial activities, improper waste management practices, and mining activities with potential threat to the environment and human health (Vidali, 2001). Mining causes environmental pollution through heavy metals accumulation (Vega et al., 2004). Revegetation of mine tailings has been the most effective method of preventing erosion and the consequent spread of contaminants to surrounding areas. However, plant growth is affected by limiting soil factors such as low pH, low fertility, high heavy metal concentration, and a small seed bank to initiate establishment (Gay and Korre, 2006). Therefore, improving soil physical and chemical properties is required for successful revegetation programs.
Nickel is a fairly rare and rather expensive metal and its main ores are nickel sulfides or nickel laterites. Nickel is rather immobile and it is enriched during the weathering process relative to its source rock, because major components like magnesium oxide (MgO) and silicon dioxide (SiO2) are leached out during weathering (Schuiling, 2013). Ultramafic rocks (rocks rich in olivine and/or serpentine) have a low Ni-content of about 0.3% (Wilson et al., 2009). Nickel is used for alloying, as a catalyst in chemical reactors, for battery making and metal plating among other uses. Its route of exposure to animals is through drinking contaminated water, atmospheric dust and bioaccumulation in plants (Lawrence et al., 2004). Several laws have been enacted to reduce these effects to humans and the environment.
The Ministry of Mines and Mining Development governs the responsibility of administering the Mines and Minerals Act (CAP 21:05) of 1997 (Government of Zimbabwe, 1997). The Act states that, ‘the rehabilitation of the surface land concerned in any mining shall be carried out by the mining authority as an integral part of the mining operations’. The Act also prescribes that in every case where vegetation, including trees, shrubs and grasses are disturbed with mining activities re-establishment should be considered. According to the Mines and Minerals Act (CAP 21:05) the reclamation of mining sites should be done once mining operations have been stopped. The Environmental Management Act (CAP 20:27) of 2002 also points the right of every citizen to a clean and safe environment and recommends the rehabilitation of disturbed environments through mining activities (Government of Zimbabwe, 2002).
2. Effects of revegetating mine tailings on soil organic carbon
2.1 Definition of soil organic carbon
Soil organic carbon refers to the difference between organic carbon inputs (vegetation, roots) and output (CO2 from microbial decomposition) (Baldock and Skjemstad, 1999). The below-ground biomass constitutes all the live and dead roots (Eggleston et al., 2006), plays an important role in the carbon cycle by transferring and storing carbon in the soil. The distribution of C down the soil profile varies with the soil type, climate, soil-plant management practices, mineral composition, topography, soil biota and the interactions between these factors (Jobbagy and Jackson, 2000; Krull et al., 2003). Changes in natural living conditions of the soil systems such as conversion of a particular land to agriculture, mining activities, settlement and plantation may result in different conditions under which SOC enters and leaves the system (Baldock and Skjemstad, 1999). The availability of vegetation on mine tailings facilitates the decomposition of SOM which in turn improves physical properties of the soil (Nelson and de Jong, 2003).
Soil organic carbon constitutes many fractions that vary in the rate of decomposition and are very heterogeneous in structure (Martens, 2000). Variations in the components of SOC results in different rates of turnover or residence times, ranging from labile to stable forms (Baldock and Skjemstad, 2000; Krull et al., 2003, Mujuru et al., 2014). The light fraction and particulate organic carbon pool are often considered the active, labile pool have a relatively fast turnover time of less than 10 years, a slow pool (mean residence times (MRTs) of approximately 25 years), and a passive, recalcitrant pool (MRTs approximately 100 to 1000 years) (Dalal and Chan, 2001).
Soil organic carbon is directly proportional to SOM content (Brunetto et al., 2006; Salehi et al., 2011). Soil organic carbon multiplied by a factor of 1.75 gives the organic matter content (Katyal and Sharma, 1991). According to Walkley and Black, (1934) SOC constitutes to about 55 – 58% of soil organic matter. Organic matter is the major source of nutrients such as nitrogen, phosphorus and potassium (Donahue et al., 1990).
2.2 Methods of determining SOC
The wet oxidation method can be used to determine SOC through acid dichromate oxidation (Kalembasa and Jenkinson, 1973), also known as the Walkley-Black method (no heating) or Heanes method. The Walkley-Black method is less accurate in determining organic C and recoveries can be as low as 56% (Skjemstad and Taylor, 1999). Determination of SOC by dry combustion or loss-on-ignition (LOI) converts all carbon in the presence of oxygen to CO2 during the heating process (Nelson and de Jong, 2003). Dry combustion is an inexpensive method that involves the heating of soil samples at high temperatures (between 300 and 6000 C for a range of 6 to 12 hours) and measuring weight loss after ignition (Brunetto et al., 2006; Konen et al., 2002). Dry combustion method also avoids chromic acid wastes (Salehi et al., 2011).
Different amounts of crop residues can result in different values of organic carbon in soils (Biswas and Mukherjee, 1994). Optimal temperatures and durations to maximize SOM combustion and minimizing inorganic carbon combustion are major challenges in determining SOC. These variables can affect SOC results and higher temperatures often drive off structural water and other inorganic constituents from soil tailings (Konen et al., 2002). Kalembasa and Jenkinson (1973) also reviewed both dry and wet oxidation methods and concluded that dry combustion methods were more accurate than the wet oxidation process. Baldock and Skjemstad (1999) also recommended the analysis of SOC by dry combustion and measurement of CO2 with an infrared detector as it is considered more accurate than the wet oxidation process.
2.3 The role of organic carbon in improving soil properties of mine tailings
Soil is an important medium in the global C cycle. Soil organic carbon plays an important role in soil biological (provision of substrate and nutrients for microbes), chemical (buffering and pH changes) and physical (stabilization of soil structure) properties of mine tailings (Jobbagy and Jackson, 2000). Soil organic carbon, nitrogen and phosphorus are considered important indicators of a quality and health soil (Donahue et al., 1990). According to Lal (2001) soil quality refers to the capacity of soil to produce economic goods and services and regulate its environment. Soil quality is an environmental buffer that helps in protecting watersheds and groundwater from agricultural and mining chemicals and municipal wastes; by sequestering carbon that helps mitigate global climate change (Reeves, 1997). However, SOC is the most important indicator of soil quality and vegetation sustainability on mine tailings (Rowell, 1994).
Changes in land cover or land use practices may affect C pools and fluxes of SOC and this often has implications on the carbon cycle and the global climatic system (IPCC, 2000). Lal, (2001) revealed that total soil C pool in forests is approximately three times the atmospheric pool and 3.8 times the vegetation pool. Lal (2004) stated that organic carbon in soils is sequestered through direct and indirect ways. The direct soil C sequestration involves inorganic chemical reactions that help to convert CO2 into soil inorganic C compounds such as calcium and magnesium carbonates. The indirect plant C sequestration occurs when plants photosynthesize atmospheric CO2 into plant biomass which may be indirectly fixed as SOC during decomposition processes.
Changes in land use and management practices influence the amount and the rate of organic C levels at regional or catchment scale (Franzluebbers, 2002). A study in Laos by Chaplot et al., (2009) revealed SOC content of 112 Mg ha-1 under forestland, 109.2 Mg ha-1 under fallow and 108.8 Mg ha-1 under continuous cultivation at 0-30 cm depth. Chiti et al., (2011) found a mean SOC content of 63.3 Mg ha-1 in rice fields and 53.1 Mg ha-1 in arable land soils at 0-30 cm depth in Italy. In another study, the C stock within the upper 10 cm of coal mine dumps ranged from 3-33 Mg ha-1 and 17-82 Mg C ha-1 in a 50 cm depth (IPCC, 2000). On six Illinois coal mine dumps spanning from 5 to 64 years in age, soil C increased by 0.1 Mg ha-1 yr-1 in the upper 10 cm and 0.3 Mg ha-1 yr-1 at the 50 cm depth (Thompson et al., 1987).
2.4 Causes of soil carbon loss from mining operations
Loss of organic C from mining operations increases the atmospheric CO2 content, at the same time reducing the fertility of the poor nutrient tailings. The SOC fractions, predominant form of C in soils of humid and sub-humid regions and soil inorganic carbon (SIC) are the predominant forms of C in arid and semi-arid regions (Powlson, 2005; Wang et al., 2001). Soil inorganic carbon (SIC) is a large pool, mainly present in the form of soil carbonates in arid and semi-arid regions (Schlesinger, 2002). The SOC fractions can release C from soil in the form of CO2 and also emit methane (CH4).
The loss of C pool through surface mining activities usually occurs by mineralization, erosion and leaching (Izaurralde et al., 2000). Changes in soil moisture, temperature regimes and reduction in the amount of biomass returned to the soil may result in the loss of C. Increase in soil temperature, increases the rate of mineralization of organic C pool (Lal, 2004). The exposed subsoil rich in calciferous materials maybe subjected to climatic factors leading to dissolution of carbonates and emission of CO2 to the atmosphere. However, the depleted SOC pool can be restored through revegetation programs (Lal, 2004). This is possible where growth and the rate of returning biomass to the soil are in excess to the mineralization capacity (Singh et al., 1996).
2.5 Bulk density an important parameter in determining SOC
Bulk density is one of the crucial factors in SOC determination and it varies with soil depth and land use system (Krull et al., 2003). Fresh mine tailings often have high bulk densities due to limited organic matter content (Maiti and Ghose, 2005), and due to relatively high heavy metal concentrations (Naumann et al., 2007). For example the density of arsenic is 5.73 g cm-3, iron 7.87 g cm-3, cobalt 8.9 g cm-3, copper 8.92 g cm-3, nickel 8.91 g cm-3 and lead as high as 11.34 g cm-3 (Duffus, 2002). Bulk density of productive natural soils generally ranges from 1.1 to 1.5 g cm-3 (Brady, 1974). In seven year old overburden dumps in India, the bulk density was as high as 1.91 Mg m-3 (Maiti and Ghose, 2005). Bulk density in the soil under a grass sward in the UK was as high as 1.8 Mg m-3 (Rimmer and Younger, 1997). The variation in bulk density of mines tailings depends on the soil texture, organic matter content and management practices (Brady, 2002; Sheoran et al., 2010). Mine soil compaction directly limits plant growth, as most species are unable to extend roots effectively through cemented surfaces. Severely compacted mine soils cannot hold enough moisture to sustain vigorous plant communities (Steinhardt et al., 1987).
2.6 Soil pH on revegetated mine tailings
Soil pH is a measure of acidity or alkalinity as determined by the amount of positively charged hydrogen (H+) ions in the soil solution (Gitt and Dollhopf, 1991), and is the most widely used indicator of mine soil quality. It is important because it influences the solubility of nutrients within the mine tailings. The pH of mine tailings can change rapidly as the rock fragments, weather and oxidize (Maiti and Ghose, 2005). Pyritic minerals (FeS2), when present, oxidized to sulfuric acid and drastically lowers the pH, while carbonate (Ca/MgCO3) bearing minerals and rocks tend to increase the pH as they weather and dissolve (Yamamoto et al., 2003). Low pH correlates closely with high dissolved metals and low metal content whereas high pH is associated with low dissolved metals and high soil metals (Barcelo and Poschenrieder, 2004). The availability of vegetation can optimize soil pH in mine tailings due to the effect of phytoremediation (Barcelo and Poschenrieder, 2003).
Boateng et al., (2012) reported an average pH range of 7.6 to 8.4 (on a water scale) at gold ex-tailing sites and a pH of 6.0 for a natural forest 16.1 km away. A pH between 7 and 8.5 indicates the solubility of calcium carbonate (Foth, and Ellis, 1997). When the soil pH drops below to 5.5, reduced plant growth occur due to metal toxicities such as aluminium or manganese, phosphorus fixation, and the population of N-fixing bacteria is reduced (Das and Maiti, 2005; Sheoran et al., 2010). A mine dump pH in the range of 6.0 to 7.5 is ideal for vegetation growth (Foth, and Ellis, 1997).
Acidic mine soils can be effectively neutralized by applying either cement kiln dust (CaO) or limestone (CaCO3). Addition of lime reduces the mobility of heavy metals in soils and increases soil pH through bio-accumulation in plants. Plants like Gravellia robusta, have been planted on acidic mine tailings as they can tolerate such living conditions (Gitt and Dollhopf, 1991). Organic amendments such as biosolids, woodchips, composted green waste or manure can moderate soil structure and pH of mine tailings. Organic amendments buffers the soil against major swings in pH by either taking up or releasing H+ into the soil solution, making the concentration of soil solution H+ more constant. The result is a stable pH close to neutral or suitable for the specific vegetation type. In addition, organic amendments on mine tailings also improve water holding capacity, cation exchange capacity, provide a slow-release fertilizer and serve as plant microbial inoculum (Tordoff et al., 2000).
3. Aboveground biomass
3.1 Contribution of trees to biomass
Tropical forests are a key component of the global carbon cycle (Malhi and Grace, 2000). Live tree biomass estimates are essential for carbon accounting, bioenergy feasibility studies, and other analyses. Trees contain about 40% of global terrestrial carbon, accounting for more than half of global gross primary productivity, and sequestering large amounts of CO2 from the atmosphere (Beer et al., 2010; Grace, 2004; Pan et al., 2011). In tropical forests worldwide, about 50% of the total carbon is stored in aboveground biomass and 50% is stored in the top 1 m of the soil (Dixon et al., 1994). An African moist tropical forest has more than three times as much carbon in aboveground biomass as in soil to 1 m depth (Djomo et al., 2011).
There are four carbon pools of terrestrial ecosystem involving biomass, namely the aboveground biomass, belowground biomass, the dead mass of litter and woody debris (Eggleston et al., 2006). The aboveground biomass of a tree constitutes the major portion of the carbon pool (Brown et al., 1989), and is the most important and abundant in terrestrial forest ecosystem (Ravindranath and Ostwald, 2008). Soil organic matter contributes secondly to carbon stocks of forests following the aboveground biomass (Lal, 2005; Kumar et al., 2006). Changes in the land use system like forest conversion, mining activities and construction have direct impact on above-ground biomass.
3.2 Methods of estimating aboveground biomass
Estimation of the accumulated biomass is crucial for assessing the productivity and sustainability of forest communities. Biomass estimation of the revegetated mine tailings enables the calculation of the amount of CO2 that can be sequestered by trees. Tree biomass can be estimated through field measurements and geographic information system (GIS) and remote sensing methods (Lu, 2006; Ravindranath and Ostwald, 2008). Field measurements of biomass can be done using destructive or non-destructive methods.
The destructive method, also known as the harvest method is the most direct way of estimating aboveground biomass and carbon stocks stored in the forest ecosystems (Gibbs et al., 2007). This method involves harvesting of tree components (tree trunk, leaves and branches) in the known area and measuring the weight after oven drying (Chung-Wang and Ceulemans, 2004; Devi and Yadava, 2009). The method is limited to a small area or small tree sample sizes. Destructive method determines the biomass accurately for a particular area though it is time and resource consuming (Montès et al., 2000). The method is also strenuous, destructive and expensive, and is not recommended for a large scale analysis (Montès et al., 2000; Ravindranath and Ostwald, 2008). Destructive method is also not applicable for degraded forests containing threatened species, and is usually used together with allometric equations for assessing biomass on a larger scale (Navár, 2009; Segura and Kanninen, 2005).
The non-destructive method estimates tree biomass without felling them and is applicable for ecosystems with rare or protected tree species where harvesting is not recommended (Chave et al., 2005). This involves either climbing the tree to measure the various tree parts (Aboal et al., 2005), or by measuring the diameter at breast height (DBH) using a diameter tape, tree height using a hypsometer, volume of the tree (Ravindranath and Ostwald, 2008), and wood density using the water-displacement method (Pyo et al., 2012). The biomasses of live or standing trees are calculated using allometric equations (Aboal et al., 2005; Hughes et al., 1999) and their reliability cannot be easily validated (Montès et al., 2000). Live aboveground biomass includes the tree stems, branches and leaves and the dead aboveground biomass comprises of leaf litter, dead standing plant parts, and fallen branches and stems.
The total aboveground biomass in a 12 year old Eucalyptus grandis plantation was 208.14 Mg ha-1 whilst an equally aged Casuarina equisetifolia plantation had 229.39 Mg ha-1. About 121.23 Mg C ha-1 was found in a 27 year old indigenous thicket dominated by Vachellia kosiensis (V. karroo) (van Rooyen et al., 2012). Neuman and Ford, (2006) found aboveground biomass of 4.72 Mg ha-1 in an offsite native soil, 6.71 Mg ha-1 in an 11 year old gold and copper treated mine tailings and 0.63 Mg ha-1 in an 11 year old gold and copper untreated mine tailings. Ilyas, (2013) reported different values for aboveground biomass in Acacia mangium Wild planted in unmined and mined areas. The carbon stocks were 246.91 Mg ha-1, 148.33 Mg ha-1, 114.90 Mg ha-1 and 77.96 Mg ha-1 for the seven year old umnined site, seven year old mined site, a five year old mined site and a three year old mined site respectively.
3.3 Surface litter and carbon accumulation in mine tailings
The soil biota plays an important role in regulating organic matter decomposition and nutrient mineralization in mine tailings (Andren et al., 1990). Plant species differ greatly in the chemical composition of their tissues which translates into great variation in the quality of surface litter to the soil. Older trees produce higher quantities of litter compared to younger ones because tree canopies enlarge with age, resulting in increased production and dropping of litter (Heal et al., 1997; Yamamoto et al., 2003). Vachellia and Senegalia species have been reported to be among the most suitable species that have been used for rehabilitating mine tailings (Jim, 2001), since they have a faster growth rates and are more tolerant. Vachellia and Senegalia species produce high quantities and quality litter, and are tolerate to a wide range of soil types and extremes soil pH (Yamamoto et al., 2003).
The decomposition of leaf litter involves a number of factors including the removal and/or consumption of tissues by leaf feeding invertebrates, leaching, and biochemical degradation by microorganisms (Wang et al., 1999). Generally, warmer temperatures with higher precipitation results in higher rates of decomposition, leading to less organic matter accumulation (Couteaux et al., 1991). Plant chemical composition, fauna and the soil community are known to influence litter and soil organic matter decomposition (Couteaux et al., 1991; Tian et al., 1992). Soil microbes influence much of the decomposition of surface litter and are the ultimate actors in the decomposition process. Soil macrofauna can make litter and the products of its physical and chemical degradation more available to soil microbes (Heal et al., 1997).
On a coal mine tailings in southern Ohio the litter layer of white pine contained 7.8 Mg C ha-1, surpassing the forest floor C beneath black locust (Robiniapseudo-acacia), yellow poplar, and white ash, and 4.1 Mg C ha-1 accumulated within forest floor beneath various forest species during a 30-year period (Vimmerstedt et al., 1989). Due to the inhibition of soil microbial and soil faunal decomposition rates, forest stands on acid spoil materials (pH < 4) support thicker forest floor layers than either neutral or calcareous spoils (Johnson and Todd, 1998). Mature mixed Southern Appalachian hardwood forests (80-100 year-old) contain 7-10 Mg C ha-1 in surface litter (Johnson and Todd, 1998). The carbon contained in the litter pool of the 12-year old Eucalyptus grandis plantation on post-mining reforestation activities on the KwaZulu-Natal coast in South Africa was found to be 13.34 Mg ha-1 and an equal aged Casuarina equisetifolia plantation had 10.78 Mg ha-1 (van Rooyen et al., 2012). However, forest species alters the rate of both SOC and forest floor C accumulation. 4. Revegetation of mine tailings and species diversity Species diversity refers to the variability among living organisms from all sources including, terrestrial, marine and other aquatic ecosystems and the ecological complexities which they are part of and this includes diversity within species, between species and of ecosystems (Bardgett, 2002; Magurran, 1988). Species diversity contributes to the regulation of carbon flux and climate control and water cycle. Plants are chemical engineers that play an important role in bioremediation, through accumulating pollutants in their bodies and degrading pollutants into smaller and non-toxic molecules. Most mine tailings disposal sites are devoid of vegetation (Mendez and Maier, 2008), and are considered environmentally harmful due to their high heavy metal concentrations (Conesa et al., 2007). Revegetation efforts including nitrogen-fixing, heavy metal and acid-tolerant species proved to be effective in mine tailings reclamation (Wang et al., 2000). Acacia saligna, or the Port Jackson willow, is very adaptable, fast growing tree native to Western Australia and has been extensively used for stabilizing soils in tailings dumps (Midgely & Turnbull, 2003). A. saligna is an invasive species with a wide range of impacts which coppices well on poor soils and calcareous sands (Midgely and Turnbull, 2003). The diversity of a plant community is greatly influenced by soil pH as it affects the availability of elements (Kabata-Pendias and Pendias, 2001). The concentration of heavy metals decreases with distance from the tailings dumps and the presence of biodiversity assumes the same trend (Jung and Thornton, 1996). Less mobile pollutants will stay inert in mine dumps while highly mobile molecules will easily contaminate the nearby communities (Warhurst, 2000). Leaf chlorosis, disturbed water balance and reduced stomatal opening are characteristic effects of toxic Ni concentrations as part of heavy metal toxicity syndrome (Clemens, 2006). Soils rich in chromium, cobalt, iron, and Ni and deficient in nutrients like calcium, molybdenum, nitrogen, phosphorus, and potassium support serpentine floras such as Helianthus exilis, Senecio clevelandii, Mimulus nudatus among other species, which are characterized by short species (Barcelo and Poschenrieder, 2004). Halophyte plants such as barley (Hordeum vulgare L.) can grow on saline soils with sodium chloride, carbonate, or sulphate; have a high osmotic pressure in cells and can accumulate high levels of salts (Brooks, 1983). Calciphilous floras grow on limestone and dolomite; include both plants, which absorb calcium and calciphilous plants, and require specific growth conditions of calcarious rocks (Brooks, 1983). Revegetation of mine tailings is the most fundamental aspect in the improvement of soil quality, assemblage of plant species and the physiological state of the disturbed ecosystem (Hao et al., 2004). Grasses are considered as a nurse crop for an early vegetation growth and are important in stabilizing soils, but may compete for nutrients, space and water with herbaceous species (Singh et al., 2002). Grasses especially C4 plants can tolerate to drought, low soil nutrients and other climatic stresses (Martins et al., 2004). Herbaceous, grasses and woody vegetation are important tools in the net removal of CO2 and climate change mitigation. A study by Mutyavaviri, (2006) at Monavale vlei in Harare, reported species richness for grasses and herbs of 60 and a diversity index (H’) of 2.83. 5. Effects of mine tailings revegetation on genetic diversity Plant populations are not homogeneous entities, and genetic diversity is as important as species diversity on reclamation sites. Wide genetic diversity enables the species to inhabit a larger number of microsites and to adapt more to environmental changes (Handel et al., 1994). Habitat destruction results in biodiversity losses: direct poisoning by mine pollutants and indirect poisoning through soil and water which affects animal, vegetation and microorganisms’ diversity (Mummey et al., 2002). Revegetated mine tailings’ dumps often exhibit a much lower (alpha) biodiversity than natural forests, due to the impact of heavy metal toxicity in mine tailings (Bradshaw, 2005). The most direct way of creating plant diversity on mine tailings is to diversify ecological community (Crowder et al., 1982; Piha et al., 1995). The indirect approach for soil nutrient enhancement involves the growing of plant species that improves the nutritional status of the soil (Haering et al., 2000, Norland, 2000). Nitrogen fixers are the most predominant and these include Senegallia polycantha, Vachellia gerradi and V. nilotica among other species (Tordoff et al., 2000). Direct-seeding may aid in the enhancement of biodiversity in poor nutrient soils. Seeding of trees and grasses in strips and intercropping them with strips of trees and grasses of the same family but from different regions promotes genetic diversity (Rosenzweig, 1995). The use of diverse plant species in introducing genetic diversity into mine tailings is crucial and appropriate for specific ecotypes that closely resemble the reclamation site (Handel et al., 1994). The diversity of plant species in marginal environments reduces the chances of plant vulnerability to pests and diseases and the community for soil flora and fauna is also promoted (Jim, 2001; Singh et al., 2002; Winterhalder, 1994). This may lead to recombination and the formation of genotypes that are better suited to the degraded habitat. Plants of local provenance and similar environments positively influence genetic diversity. Millar and Libby (1989) revealed that lack of local native stock of desired species can result in the build up of a new landrace from a mixture of distant populations. Direct seeding of wild species, transplanting of young individuals from natural sites and applying layers of litter are the best approaches for genetic diversity (Winterhalder, 1994). 6. Role of soil macro-fauna in improving mine tailings Soil fauna are a diverse community of soil-dwelling animals. Faunal species are functionally classified according to body width; microfauna are < 100 micrometers (μm) wide, mesofauna 100 μm to 2 mm wide, and macrofauna are > 2 mm wide (Coleman and Wall, 2007; Ghose, 2005). Macrofauna species common in soils include microscopic hair-like worms called nematodes, miniature earthworm-like animals called pot worms or enchytraeids, snails, slugs, springtails, insect larva, beetles, ants, spiders and earthworms (Sugiyarto, 2000). Mesofauna is represented by anthropods such as mites, collembola and enchytraeids, insects and their larvae, round and annelid worms, and some species of terrestrial crustaceans and mollusks (Gongalsky, 2000). Soil microorganisms are those organisms that can only be seen with the aid of a microscope and these include bacteria, actinomycetes, fungi, algae and protozoa (Coleman and Wall, 2007).
Soil organisms are crucial in the decomposition, soil formation, nutrient cycling and maintaining of soil structural conditions and moderate many physical and chemical processes in mine tailings (Coleman and Wall, 2007). Earthworms, centipedes, ants, etc. are important in the translocation of plant residues (Rai, 2002). Dead root, litter and leaf drop, and the bodies of soil animals such as insects and worms are the primary sources of organic matter in the soil (Daniels and Haering, 2006). Soil fauna are important elements in the re-establishment of mine tailings’ ecosystem.
The consumption of plant residues together with the burrowing activity mixes organic matter with soil and enhances soil fertility (Coleman and Wall, 2007). Soil tunneling activity improves gas exchange, water infiltration and root proliferation. Soil material excreted by earthworms (earthworm cast) is high in plant available nutrients and helps in improving soil structure (Allen et al., 2001). Macro fauna helps in the degradation of agrochemicals and pollutants in both agricultural land and mine tailings. Soil fauna is strongly affected by changes in moisture and temperature regimes beyond the normal parameters of seasonal patterns as well as the availability of vegetation cover (Dangerfield et al., 1998). The higher the vegetation cover, the more the presence of macro fauna. Plant litter and root exudes energy and nutrients for soil biota and support channels of soil and rhizosphere food webs (Coleman et al., 2004).
A 20-year old nickel dump had species richness of 28 mite (Acari) whilst a copper tailings had a higher diversity (H’ = 1.79), a 40-year old tailings had species richness of 20 and a diversity index (H) of 1.31 whereas an eight year old tailings had a species richness of 15 and a diversity index of 1.21 (John et al., 2002). In a study by Mutyavaviri, (2006) at Monavale vlei in Harare, species richness for insects was found to be 27 and a diversity index of 2.54.
Re-vegetating mine tailings has been seen as one of the best option of making these areas useful for use by future generations. Soil fertility will be restored and these areas can be used for farming. Genetic biodiversity is also improved with new plant species being introduced through succession. Agroforestry species mine all leached nutrients through deep capture system and return these back on surface increasing microbial population in the area.