Environmental Pollution (Series A) 33 (1984) 121-131
Chemical and Microbiological Changes in Soil Following Exposure to Heavy Atmospheric Pollution K. Killham* & M. Wainwright Department of Microbiology, University of Sheffield, Sheffield S10 2TN, Great Britain ABSTRACT Soils exposed to atmospheric pollution from a coking works were more acidic and contained higher concentrations of S-oxides and higher numbers of S-oxidising micro-organisms than did similar, but unpollutedt, soils. Transfer of unpolluted soil to the polluted site resulted in an increase in S-oxidising micro-organisms and S-oxides in these soils, and to decreases in soil pH and the number of heterotrophic bacteria. Transfer of polluted soil to the unpolluted site did not, however, result in a marked increase in soil pH, or in a change to more normal concentrations of S-oxides and numbers of S-oxidising micro-organisms. The fact that the concentration of S-oxides in the polluted soil did not fall when removed from exposure to the effluent indicates that internal microbial oxidation accounts to some extent for the higher concentration of Soxides and lower pH found in polluted soil. The results suggest that without treatment, the brown earth soils studied here would not rapidly recover once the source of the pollution had been removed, either by effluent clean-up or plant closure. INTRODUCTION While the effects o f a t m o s p h e r i c pollution f r o m coal-fired power stations a n d smelters on soils a n d vegetation have been extensively studied ( F r e e d m a n & H u t c h i n s o n , 1980; W a i n w r i g h t , 1980a), few c o m p a r a b l e * Present address: College of Natural Resources, Department of Plant and Soil Biology, University of California, Berkeley, California 94720, USA. t The term unpolluted site is relative as this site receives SO z but not coke oven derived gases or particulates. 121 Environ. Pollut. Ser. A. 0143-1471/84/$03.00 © Elsevier Applied Science Publishers Ltd, England, 1984. Printed in Great Britain
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K. Killham, M. Wainwright
reports exist on the effects ofcoking and smokeless fuel plant effluents. At modern installations pollution from these last two sources is usually stringently controlled, but many old and inefficient plants still operate (Pearce, 1981) leading to localised but heavy pollution. Smokeless fuel and coking effluents are often rich in particulates as well as SO 2, and as a result vegetation growing close to the effluent source is frequently covered with a thin layer of soot (Wainwright, 1978a). Our recent studies have shown that soils exposed to pollution from coking works are richer in Soxides and S-oxidising micro-organisms, and have a lower pH than similar, but unpolluted, soils (Wainwright, 1980b; Killham & Wainwright, 1981). These trends, which tend to be localised around the source of pollution, are accentuated beneath tree canopies, due to enhanced particulate-capture and wash-off from the surface of vegetation. Most workers consider that direct input from the atmosphere accounts in large part for the reduction of soil pH and high S content of polluted soils, but we have recently demonstrated (Killham & Wainwright, 1982) that micro-organisms can release S ions from the atmospheric pollution deposits emitted from coking works, so that some of these changes may result from microbial transformations. The aim of the work reported here was (a) to compare the initial chemical and microbial properties of soil 0-5 km downwind of a coking works with those of a similar, but unpolluted, soil; (b) to monitor the changes which occurred over 2.5 years, when an unpolluted soil is transferred to a polluted site, and vice versa. These experiments were designed to show how a soil responds to a new pollution load and the speed of natural recovery shown by a polluted soil to an improvement in effluent quality or to plant closure. METHODS Samples of the top 10 cm of soil were removed from beneath the canopy of sycamore Acer pseudoplatanus L. at both the polluted and unpolluted sites. Soils at both sites were brown earths showing signs of podsolization. Climate and vegetation (mixed deciduous) at the sites were similar (Table 1), but the polluted soils have been exposed to coking pollution for over 50 years, the present SO 2 levels averaging 150#gm -3, but occasionally exceeding 200 #gm- 3. The soils were packed into open-ended plastic tubes (10 cm length x 6.5 cm diameter), which were then buried flush with the surface of soil
123
Chemical changes in polluted soil TABLE 1 Chemical Properties of Soils and Climatic Data (1979) for the Two Sites
Soil pH
Total soil C (9/o) Total soil N (9/o) Mean daily maximum temperature (°C) Mean daily minimum temperature (°C)
Annual rainfall (mm) Number of days on which rain fell Number of days with snow lying on the ground
Polluted site
Unpolluted site
3.3 14.6 0.93 12.7 6.2 830 181 23
4-9 9"8 0.82 12.9 6.3 805 173 27
below sycamore canopies. Soils were packed as uniformly as possible, so as to achieve soil densities similar to those found in the undisturbed environments. The indigenous litter layer was then replaced. At each site, samples of both native soil and soil from the other site were buried (thirty tubes for each soil type) and six tubes for each soil type were removed at 6month intervals at both sites over a period of 2.5 years. On removal from the field, the soil from each tube was mixed thoroughly and sieved ( < 2 mm) prior to chemical and microbial analysis. Results are expressed as means from six soil samples, three replicates per sample. Student's t-test was used to test for significant differences. Viable counts of micro-organisms were determined using the plate count technique. Soil (1 g) was shaken for 15 min in sterile Ringer's solution (1/4 strength, 100 ml) on a reciprocal shaker (100 throws min- 1). Aliquots (0.1 ml) of the desired dilution were then spread onto the surface of suitable media. Numbers of chemoautotrophic, S-oxidising bacteria and S-oxidising fungi were estimated using the media of Wieringa (1966) and Wainwright (1978b), respectively. Clearing of the opaque medium overlay was the criterion used to indicate presumptive S-oxidation. 'Total' counts of heterotrophic bacteria and fungi were determined using Plate Count agar (Lab M) and Czapek Dox agar (Oxoid), plus 50 mg litre-1 actidione and streptomycin, respectively. Results are expressed as means from six soil samples, fifteen plates per sample. SO~- was determined turbidimetrically (Hesse, 1971) and $20 ~- and $40 ~ - by colorimetry (Nor & Tabatabai, 1976). To extract S-ions from the soils, samples (1 g, sieved to < 2 m m ) were shaken for 15 min in 10ml
124
K. Killham, M. Wainwright
LiC12 (0"IM), and then filtered through Whatman No. 1 filter paper. Organic C and total N were determined by the Walkley & Black (1934) and macro-Kjeldahl methods (Hesse, 1971), respectively; cation exchange capacity.and base saturation by Brown's (1943) method, while soil pH was determined using a glass electrode and a 1 : 1 soil :water slurry, shaken for 15 min.
RESULTS AND DISCUSSION Soils at the polluted site initially had higher levels of organic C and total N than did unpolluted soils (Fig. 1), but unpolluted soils rapidly accumulated organic C (and, to a lesser extent, total N) when left at the polluted site for 2.5 years (Fig. 1). Neither the organic C nor N contents of the polluted soils, on the other hand, changed significantly when transferred to the unpolluted site (Fig. 1). These changes presumably result from the inhibitory effects of the coking pollution on litter degradation. It appears, therefore, that the turnover of organic matter in the brown earth studied is impaired by exposure to heavy atmospheric pollution. Previous studies by us (Killham & Wainwright, 1981) have also shown that sycamore-litter degradation is reduced at the polluted site, due apparently to a reduction in primary leaf-shredding by soil invertebrates. The polluted soil had a significantly lower pH than unpolluted soil, although the cation exchange (CEC) capacities of both soils were the same at 19 meq per 100 g of soil. The C EC of the unpolluted soil was composed of 33.3 ~ exchangeable bases, however, compared with only 11.7 ~ for the polluted soil. This suggests that both soils are essentially similar in their clay and organic matter contents but that the unpolluted soil has had much of its interchangeable bases (Ca2+; Mg2÷; K ÷ and Na ÷, etc.) replaced by H ÷ or H3 O÷, due to infiltrating rain, or to H ÷ produced during microbial or abiotic S-oxidation. Acidification of the unpolluted soil occurred when it was transferred to the polluted site, the pH decreasing by 1.0 unit from 4.9 to 3.9 over 2.5 years (Fig. 1). Most of this pH change occurred following an initial period of stability when the pH of the exposed unpolluted soil remained unchanged over the first 12 months. The unpolluted soil apparently had an inherent buffering capacity which tended to resist marked falls in pH. Wiklander (1973/74) suggested that the acidification of soil resulting from
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126
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infiltrating acid precipitation could be viewed as a three-stage process: (i) easily soluble salts are leached or translocated to deeper horizons by percolating excess water; (ii) more sparingly soluble compounds (e.g. CaCO3) are then gradually brought'into solution, and leaching gradually reduces the soil solution concentration of bases, which enables ion exchange, with H ÷ and H 3 0 + replacing the base cations Ca 2 ÷ ; Mg 2 ÷ ; K ÷ and Na ÷ and (iii) soil pH and base saturation decreases markedly and the soil becomes less fertile. The unpolluted soil appears to have undergone acidification in this way, with a major pH drop occurring only after readily leached salts were removed. Increase in the pH of the polluted soil on transfer to the unpolluted site occurred much more slowly than the above acidification of the unpolluted soil (Fig. 1) with an increase of only 0.3 of a pH unit occurring over 2.5 years. This indicates that if the coking works were closed, or the quality of the effluent were improved, the pH of soils adjacent to the works would not recover rapidly to its pre-pollution level (unless aided by liming). The concentration of SO 2- in the polluted soil was 1.7 times greater than that of the unpolluted (Fig. 2), S 2032 - and $4062 - were absent from the unpolluted soil but present in appreciable amounts in the polluted soil (Fig. 2). The presence of these latter ions suggests that microbial Soxidation is occurring. $2032 - and S,O~ - appeared in the unpolluted soil when it was transferred to the polluted site (Fig. 2), but little change in the concentration of these ions occurred in the polluted soil when it was left at the unpolluted site (Fig. 2). This is probably because there was sufficient reduced S available to micro-organisms in the soil to allow for the production of S-oxides to continue. Had $202- and $4062- in the polluted soil been derived from external sources one would have expected their concentration to fall off dramatically when soil was transferred to the unpolluted site (which did not receive these external inputs); the fact that they did not adds weight to the view that microbial oxidation accounts for the presence of these ions and some of the SO 2- found in the polluted soils. Counts of S-oxidising micro-organisms were also higher in polluted than unpolluted soils (Fig. 3) and numbers of both thiobacilli and Soxidising fungi increased in the unpolluted soil on exposure to pollution (Fig. 3). Previous reports from this laboratory have shown that the numbers of S-oxidising micro-organisms are higher in soils exposed to air pollution from a variety of sources, compared with similar but unpolluted soils (Wainwright, 1978a,b), and independent verification of this trend
Chemical changes in polluted soil
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K. Killham, M. Wainwright
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130
K. Killham, M. Wainwright
has recently been reported for air-polluted soils in Czechoslovakia (Lettl et al., 1981). The polluted soil contained less than half the heterotrophic bacteria found in the unpolluted soil (Fig. 4), and a marked reduction in numbers in the unpolluted soil occurred when it was transferred to the polluted site (Fig. 4). These changes probably reflect the general susceptibility of heterotrophic bacteria to acidity (Gray & Williams, 1971). The number of fungi determined by the plate count method was not, however, reduced by exposure to pollution (Fig. 4). In a previous report (Wainwright, 1980b), it was shown that a 1-year exposure to the air pollution studied here had little effect on soil microbial activity. The present studies confirm that this exposure period is too short for the influence of this pollution on soil properties to become apparent, since most of the effects on soil micro-organisms and soil chemistry appeared only after a 12-18 month exposure. In conclusion, the chemistry and microbiology of a brown earth soil was fundamentally changed following a 2-5-year exposure to pollution from a coking works. Increases in the concentration of 52O2- and 54O2- and, to a lesser extent, SO]--, together with increases in the number of S-oxidising microorganisms, suggest that microbial oxidation is occurring in these soils (Wainwright, 1979). Polluted soils showed no sign of recovery over the 30-month incubation period, when transferred to an unpolluted site. It is clear, therefore, that the closure ofcoking works would not automatically result in a rapid recovery of nearby soils to their pre-pollution pH, and Sion content. ACKNOWLEDGEMENT Financial assistance from NERC for a PhD studentship for K.K. is gratefully acknowledged. REFERENCES Brown, I. C. (1943). A rapid method of determining exchangeable hydrogen and total exchangeable bases of soils. Soil Sci., 56, 353-7. Freedman, B. & Hutchinson, T. C. (1980). Pollutant inputs from the atmosphere and accumulation in soils and vegetation near a nickel-copper smelter at Sudbury, Ontario, Canada. Can. J. Bot., 58, 108-32.
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Gray, T. R. G. & Williams, S. T. (1971). Soil micro-organisms, London, Longman. Hesse, P. R. (1971). A textbook of soil chemical analysis, London, J. Murray. Killham, K. & Wainwright, M. (1981). Deciduous leaf litter and cellulose decomposition in soil exposed to heavy atmospheric pollution. Environ. Pollut. (Series A), 26, 69-78. Killham, K. & Wainwright, M. (1982). Microbial release of sulphur ions from atmospheric pollution deposits. J. appl. Ecol., 18, 889-96. Lettl, A., Langkramer, O., Lochman, V. & Jaks, M. (1981). Effect of industrial emissions with high sulphur dioxide content on thiobacilli and oxidative activity of spruce torest soils towards inorganic sulphur compounds. Folia Microbiol., 26, 151-7. Nor, Y. M. & Tabatabai, M. A. (1976). Extraction and colorimetric determination of thiosulfate and tetrathionate in soils. Soil Sci., 122, 171-5. Pearce, F. (1981). Fear of dole keeps Britain's dirtiest factory open. New Scient., 90, 745. Wainwright, M. (1978a). Distribution of sulphur oxidation products in soils and on Acer pseudoplatanus L. growing close to sources of atmospheric pollution. Environ. Pollut., 17, 153-60. Wainwright, M. (1978 b). A modified sulphur medium for the isolation of sulphur oxidizing fungi. PI. Soil., 49, 191-3. Wainwright, M. (1979). Microbial S-oxidation in soils exposed to heavy atmospheric pollution. Soil Biol. Biochem., 11, 95-8. Wainwright, M. (1980a). Man-made emissions of sulphur and the soil. Int. J. environ. Stud., 14, 279-88. Wainwright, M. (1980b). Effect of exposure to atmospheric pollution on microbial activity in soil. Pl. Soil, 55, 199-204. Walkley, A. & Black, I. A. (1934). An examination of the Degtjareft method for determining soil organic matter and a proposed modification of the chronic acid titration method. Soil Sci., 37, 29-38. Wieringa, K. T. (1966). Solid media with elemental sulphur for detection of Soxidizing microbes. Anton. van Leeuwenhoek, 32, 183-6. Wiklander, L. (1973/74). The acidification of soil by acid precipitation. Grundforbattring, 26, 155-64.