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Effects of Copper Toxicity on Soil Microbial Biomass



Ayman M. El-Ghamry, Abid Subhani, Wael Mohd, Huang Changyong and Xie Zhengmiao
 
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ABSTRACT

The wide utilization of copper in agriculture and industrial processes in addition to many other uses associated with human activities causes both point and non-point source pollution of the environment with copper. Although copper is relatively non-toxic to mammals and essential micronutrient necessary for a wide variety of prokaryotic and eukaryotic metabolic activities, is toxic to all organisms if present in elevated concentrations. Copper pollution in agricultural posing soils can adversely affect the living part of soil organic matter serious threat to sustained food and fiber production.

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  How to cite this article:

Ayman M. El-Ghamry, Abid Subhani, Wael Mohd, Huang Changyong and Xie Zhengmiao, 2000. Effects of Copper Toxicity on Soil Microbial Biomass. Pakistan Journal of Biological Sciences, 3: 907-910.

DOI: 10.3923/pjbs.2000.907.910

URL: https://scialert.net/abstract/?doi=pjbs.2000.907.910
 

Introduction

Previous investigations carried out on the effects of heavy metals on soil microorganisms invariably indicated the adverse effects of these elements on the soil microbial activity, number and population (Olson and Thornton, 1982; Duxbury and Bicknell, 1983; Maliszewska et al., 1985; Hiroki, 1992; Chander and Brookes, 1991, 1993; Hattori, 1992; Huysman et al., 1994; Doelman and Haanstra, 1984; Aoyama and Nagumo, 1996; 1997). For example excessive heavy metal concentration in the soil has been reported to cause a decrease in microbial population (Hicks et al., 1990; McGrath et al., 1995), changes in populations structure (Chaudri et al., 1993; Bardgett et al., 1994; Huysman et al., 1994) and physiological activity (Bitton and Dutka, 1986; Cotrufo et al., 1995; Valsecchi et al., 1995). The role of microbial biomass in soil fertility by serving as a reservoir of plant nutrients (esp. N, P and S) and as catalyst in the cycling of C, P and S (Smith and Paul, 1990; Lee and Pankhurst, 1992) make microbial biomass indispensable in soil nutrient flux and bioavailability. Toxicity of heavy metals in plants and animals have been well established but little is known about their adverse effects on soil microbial biomass (Baath, 1989; McGrath et al., 1995). Of these metals, copper is extensively used in agriculture in form of metal containing fertilizers, bactericides, fungicides, algicides and in form of metal contaminated composts, sewage sludges, as well as feed additive in antibiotics, drugs, growth promoters etc, (Adriano, 1986). Its abundance in the earth crust (24-55 μg g–1) and soil (20-30 μg g–1) (Alloway, 1990) makes it a pollution problem in most agricultural soils. Copper is highly toxic to microorganisms if present in excess concentration (Flemming and Trevors, 1989; Maliszewska et al., 1985; Alloway, 1990) which consequently cause change in soil biological equilibrium with adverse effect on both soil fertility, plant development and yield (Huisingh, 1974). A recent study by He et al. (1997) indicated a highly significant correlation between the biomass-C, N and the N uptake by plants indicating that microbial biomass-C and N are not only important soil fertility parameters but also important indices of soil sustainability. This paper therefore reviews the effect of copper toxicity on soil microbial biomass with particular attention to microbial biomass-C, N and enzyme activity.

Effect of copper on soil microbial biomass: Soil microbial biomass consists of many species of bacteria and fungi together with larger soil organisms such as yeast, algae and protozoa. However, inspite of the specie specific role in soil, microbial biomass is often considered as a single compartment (Brookes et al., 1985). Previous studies carried out on microbial biomass in metal polluted soils indicated decrease in soil biomass e.g., (Brookes and McGrath, 1984; Brookes et al., 1986; Zibilske and Wagner, 1982). Maliszewska et al. (1985) put the relative toxicity of metal compounds on soil microorganisms in the following order: Hg (metallic)> CuSO4> HgCl2> Hg2Cl2> HgO> As2O3> ZnSO4> NaHAsO4> Pb (NO3)2.

Copper is known to be very toxic to microorganisms in the free ionic form especially Cu2+ and CuOH+ (Zevenhuizen et al., 1979). The addition of copper to soil was reported to significantly decrease the amount of microbial biomass and exert a pronounce toxic effect on the decrease size of the biomass compared to certain metals such as Pb and As when compared on a molarity basis (Aoyama and Nagumo, 1997). Also he concluded that in apple orchard soils with heavy metal accumulation Cu was the primary factor affecting microbial biomass. Knight et al. (1997) conducted a study of the effect of Cd, Cu and Zn on microbial biomass and reported that only Cu showed a notable reduction in microbial biomass and reduced the metabolic potential of soil microbial population. In general bacteria are more sensitive to Cu than fungi and most Cu resistant bacteria are oxidase positive and 50% of the strains are pigmented (Huysman et al., 1994). Low doses of Cu (10-100 μg g–1) added to sandy and alluvial soil showed a marked harmful influence on the growth of bacteria, actinomycetes, cellulytic, microflora and nitrifiers in both soils (Maliszewska et al., 1985).

Kobus and Kabata-Pendias (1977) and Wilson (1977) had earlier observed inhibition of growth of azobacter and nitrifiers by the presence of high concentration of Cu in soil. Hattori (1992) also reported a marked decrease in soil bacterial colonies by addition of 10 μmol g–1 soil Cu to Gley and Andosol soil. He further observed that addition of 20 μmol L–1 of Cu decreased the number of colonies to one tenth of the control. Hemida et al. (1997) found that application of CuSO4 to clay soil for one week resulted in the reduction of bacterial count (p<0.05) with 200 and 2000 μg g–1 soil but same doses enhanced bacterial count in sandy soil one and four weeks after application although the count was only significant after four weeks of application.

The addition of Cu to a paddy soil, up to 1600 μg g–1 soil, was reported to increase fungal population size, proportional to the Cu concentration added, yet decreased fungal diversity by 40%. (Hiroki et al., 1985). The number of fungi in gley soil was shown to reduce with Cu increase by more than 100 times compared to control (Hattori, 1992). Similar fungal reduction in excess Cu was earlier observed (Schnurer, 1989). The increase in fungal population was related to the inhibition of CO2 evolution from soil (Hattori, 1992). Hiroki et al. (1985) had reported a positive correlation coefficient between the Cu content and the number of fungal colonies in the soil and that Cu tolerant fungi especially the genus penicillium to be dominant in Cu polluted soils. However, a contradiction was reported by Maliszewska et al. (1985) where growth of fungi was found to be stimulated by the addition of Cu in soils with high doses (5000-10,000 μg g–1 Cu). (Hemida et al., 1997). The authors observed a non significant effect on thermophillic and thermotolerant fungi in clay soil and a significant reduction in sandy soil one week after application but three weeks later the count significantly increased. They reported the count of actinomyces isolated in the two soils to be very low or not measurable during the first four weeks of application.

Zibilske and Wagner (1982) on the contrary reported that penicillium and paecilomyces decreased with increase Cu concentration in soil while Trichoderma, Rhizopus and Asprgillus increased. However, Hiroki et al. (1985) argued that it is quite probable that penicillium is the dominant genus in copper polluted soils because in their study most 1000 μg Cu ml–1 tolerant fungi were penicillium and the growth or sporulation of some of the penicillium isolated from the soil were stimulated in medium containing Cu and about 65% of all the isolates at high Cu concentrations were penicillium. Chander and Brookes (1993) observed that Cmic/Corg in soils contaminated with higher rates of Cu was less than half of that in the soils which received no sludge, uncontaminated sludge or sludge contaminated with lower rates of metals. The decline in Cmic/Corg in metal contaminated soils occurs mainly due to reduction in the rate of carbon mineralization that in turn results in the accumulation of organic matter in soil. Chander and Brookes (1991) have reported reduced rates of organic matter decomposition in metal contaminated soils and they found that Cu at about 2.5 times the permitted metal limits caused an increased accumulation of organic carbon, about 30% in sandy loam and 13% in silty clay loam soils. Witter et al. (1993) found about 60% reduction in the Nmic in soils collected from metal contaminated sludge-treated plots. The metal concentration at which this reductions occurred was 125 μg g–1 soil Cu. Also Bardgett and Saggar (1994) reported a significant decrease in the Nmic along a gradient of increasing Cu concentrations in a pasture soil.

Effect of copper on microbial enzyme activity: Microbial enzyme activity has also been previously studied in relation to heavy metal pollution in soil (Mathur et al., 1980; Nordgren et al., 1986). The activities of dehydrogenase (Wilke, 1988; Schnurer, 1989), phosphatase (Speir et al., 1992) and urease (Hemida et al., 1997) were found to decrease in a wide range of Cu polluted soils and could be used as indicators of the toxic effects of Cu on soil microbial populations. However, due to highly significant correlation between the microbial biomass and the dehydrogenase activity (Wilke, 1988), C-mineralization and dehydrogenase activity (Serra-Wittling et al., 1995) and 0.1M CaCl2-extractable Cu and dehydrogenase activity, the measurement of dehydrogenase activity was regarded as the most sensitive parameter for the detection of harmful effects of Cu on the soil microflora (Wilke, 1987,1988; Aoyama and Itaya, 1995). Several workers indicated the effect of Cu contamination on soil microbial enzymes, for example: Tyler (1974). Tyler (1981) showed lower enzyme activity with low levels of Cu concentration. Mathur et al. (1980) studied several hydrolytic enzymes in relation to Cu content of cupriferous bogs and found that although differences existed in the degree of inhibition, most enzymes were affected in a similar way. They also noticed that enzyme activities were all found to be lower in soil containing higher Cu. Yet in another study carried out by Hemida et al. (1997) copper was found to reduce urease activity in soils at the rate of 200 μg g–1soil in both clay and sandy soil. The authors also observed similar effect on nitrate reductase although much weaker than that of urease. However, they found a marginal effect on amidase activity, which was only significantly affected at high application rate of 2000 μg C g–1 soil. A somewhat similar observation had been made by Frankenberger and Tabatabai (1981) who reported that metal ion does not strongly inhibit amidase activity. High concentration of Cu was observed to markedly inhibit the activity of dehydrogenase in sandy and alluvial soils (Maliszewska et al., 1985; Chander and Brooks 1991; Chander et al., 1995). It was found a significant decline in the dehydrogenase activity in soils amended with metal-enriched sludges (Chander et al., 1995). Further study by Chander and Brookes (1991) indicated that the main reason of decreased dehydrogenase activity was the abiological reaction between TPF and Cu, rather than a biological effect. All these observation agreed with the early study conducted by Tyler (1974) who observed general inhibition of soil enzymes by excess Cu contamination.

Effect of copper on microbial activity.: Measurement of soil respiration rate was largely used over the years in relation to heavy metal pollution (Baath, 1989). Most heavy metals pollution exhibited little effect on CO2 evolution at low levels of contamination, but with higher doses, a significant decrease in soil respiration rate was reported (Capone et al., 1983; Nordgren et al., 1983; Speir et al., 1992; Aoyama et al., 1993; Cotrufo et al., 1995).

However, there were a few exceptions that reported increase in the respiration rate due to metal addition to soils (Bond et al., 1976; Capone et al., 1983). These differences suggested that measurement of soil respiration alone could not be efficient in determining the toxic effects of metals on the size and activity of the soil microbial biomass (McGrath et al., 1995). However, soil respiration rate was easy to be measured and appears to be a sensitive measurement with which to detect heavy metal pollution (Baath, 1989). The standardization of soil water content minimized the variability among the samples and can detect changes in respiration rate at lower contamination levels than soil samples with field moisture content (Tyler, 1974). Combining microbial activity and the population measurement, such biomass specific respiration appears to provide a more sensitive indication of the soil pollution by heavy metals (Brookes and McGrath, 1987; Brookes, 1995). Fliebach et al. (1994) measured respiration in soil of low and high metal sludge and observed that soil respiration especially the respiration per unit biomass (specific respiration) increased with the increasing heavy metal concentrations in soil. Recent study (Bardgett et al., 1994) measured the specific respiration (respiration per unit biomass) in the top 5 cm of a pasture soil and reported a significant increase in the ratio of respired C to biomass C along a gradient of increasing Cu concentration. In addition to other reports that indicated the negative effect of high doses of copper on CO2 evolution (Jackson and Watson, 1977; Mathur et al., 1980; Nordgren et al., 1986). Mathur et al. (1980) reported in a muck (sapric) sample containing 2922 μg g–1 of total Cu lost carbon through aerobic soil respiration at half the rate of a muck sample containing 1159 μg g–1 Cu. Similarly the soil respiration rate of the peat (hemic) sample containing 797 μg g–1 Cu was half that of the peat sample with 408 μg g–1 Cu. Hattori (1992) found that copper exerted an inhibitory effect of CO2 evolution from the Gley soil and less conspicuous effect in light colored Andisol. Earlier studies have also indicated the effect of heavy metals on N-mineralization in the field (Brooks et al., 1986) and N-mineralization had been indicted to be a very sensitive measurement of metal pollution (Tyler, 1981). Premi and Cornfield (1969) observed that a relatively large doses of Cu stimulated the activity of nitrification and ammonification in soils with 100 μg g–1 Cu and a distinct inhibition of this process with 10000 μg g–1 Cu. Brookes et al. (1986) found a 50% decrease in N2-fixation by green algae in sludge amended soil at a very low level of Cu contamination (15 μg g–1 Cu).

Copper was also observed to inhibit N2-fixation by free living heterotrophic bacteria in the soil and reduced that of cyanobacteria by 50% at the concentration of 37 μg g–1 Cu. It reduced that of Rhisobium leguminosarim by several orders of magnitude at soil Cu concentration of 27-48 μg g–1 Cu (McGrath et al., 1988). Several other reports indicated similar negative effect of Cu on N2-fixation (Porter and Sheridan, 1981;McGrath et al., 1988).

Conclusion

Most studies indicated the toxicity of excess Cu to size of soil microbial biomass, their activity and enzymes. The few contradictions may be attributed to different properties of soils, experimental methods, conditions etc. However, the effect of Cu contamination on soil microbial biomass seriously affects the recycling of nutrients in the soil and thus plant growth and yield. The combined effect of Cu interaction with other metals, microbial adaptation and subsequent tolerant to Cu by microorganisms, different soils organic matter content etc. suggest the need for a well plan standard study of the effect of copper toxicity on soil microbial biomass.

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