Lake Sediments as Indicators of Heavy-Metal Pollution Ulrich F t r s t n e r L a b o r a t o r i u m ftir Sedimentforschung, Universitfit Heidelberg, Federal Republic of G e r m a n y
Heavy metals are one of the most toxic forms of environmental pollutants, constituting a threat both to aquatic life and the quality of drinking water. By analyzing lake sediments, it is possible to determine the provenance, distribution, extent, and also the possible hazards of metal contamination. Sedimentary cores, in particular, provide the means for evaluating the different influences from natural and civilizationaI sources; they represent a historical record of the metal accumulations which have taken place during the past decades as a result of population growth and industrial development.
The significance of sediments in assessing the conditions in aquatic systems [1] is best illustrated by examples f r o m the lacustrine environment. This is a result of the fact that freshwater lakes have formed the center for important cultural developments ever since the earliest days of civilization. As a consequence of increased population and industrialization densities, the threat of pollution has become most acute in such areas [2].
Interferences: Geochemical Background and Man's Impact Environmental as well as explorational geochemistry requires a knowledge of the naturally occurring metals in order to evaluate m a n ' s environmental impact: " B o t h the exploration and environmental geochemist can be looking for the same type of areas, those with high metal concentrations, but obviously from a different m o t i v a t i o n " [3]. This is especially valid for those regions where lake sediment geochemistry is used as a guide to mineralization. A center of recent activities in this respect is Canada, which has a greater area than any other country covered by freshwater lakes [4, 5]. As early as 1956, Schmidt [6] evidenced anomalous metal distribution patterns in lake sediments bordering on areas of mineralization in New Brunswick and Quebec. F r o m studies in Saskatchewan, Arnold [7] and D y c k [8] inferred that at low Naturwissenschaften 63, 465-470 (1976)
sampling density, lake sediments reveal outline metalliferous areas as accurately as stream sediments. Lake sediment geochemistry, at the reconnaissance level, was first described by Allan [9] in a survey covering an area of 3,800 k m 2 in the Coppermine region. Since 1970, thousands of mountain lakes (Appalachia, Cordillera), Arctic lakes, prairie lakes and, in particular, shield lakes have been analyzed both for geochemical exploration and environmental management [4]. The latter aspect of pollution monitoring from lake sediments attracted much interest when mineral exploration was followed by large-scale mining and processing activities in certain areas. As shown by examples from N o r t h America, Australia, and South Africa, these operations affect lake-sediment composition in four main ways [3]: (1) tailings introduced as solids into drainage systems, (2) leachates from onshore tailing piles [10-14], (3) effluents from mines or mining plants [15], and (4) airborne particulates from crushers and smelters [16, 17]. However, the different sources of pollution listed above are generally interrelated, often rendering it impracticable to distinguish between them by means of bulk sediment analysis. The exploration-environmentalgeochemistry"overlap" in lake sediments was exemplarily investigated by Allan [3] in the large nickel-mining area at Sudbury on 26 different examples. Most of the lakes studied had accumulated less than 10 cm of sediment within the last hundred years. Thus, samples taken from below this depth reflect the natural levels which occurred prior to ore exploitation, whereas samples from the 10-cm level indicate the extent of subsequent mining activities. In Kelley Lake, adjacent to the Coppercliff smelters, the nickel content, even at levels down to 15 cm of the sediment, was elevated to such an extent that "it may prove economicallyfeasible to reclaim this or other lakes for their metal content" [3]. An example of the interface of lithologie influences and man-made pollution effects with regard to metal concentrations in lake sediments is described by Wittmann and Ftrstner [18]: The Hartbeespoort Dam, which represents one of the most highly eutrophicated impoundments of South Africa [19], is situated 40 km to the -west of Pretoria between the Daspoort Hills on the southern side and the Magaliesberg mountains constituting the northern boundary. From the western catchment,area, the dam is fed by the Magalies and Skeerpoort rivers, which drain a predominantly agricultural basin. In contrast, the eastern Crocodile river inflow stems from the highly industrialized Kempton Park-Isando area, north of Johannesburg [21]. In Table 1 the heavy-metal contents of characteristic sediment samples from the western (A) and the southeastern (B) zones and
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Table 1. Comparison of metal content in sediments from the western (A) and southeastern (B) zones of the Hartbeespoort Dam and from the corresponding inflow rivers (C, D) Cr Quotient B/A Quotient D/C Enrichment mechanism
Ni
Co
0.56 0.57 0.63 0.52 0.46 0.56 lithologic influences
Fe
Cu
Mn
Hg
Pb
Cd
Zn
0.66 0.63
0.91 0.92
0.97 0.94
1.55 2.48 2.50 4,70 2.25 1.87 man-made pollution
2.55 2.00
Table 2. Distribution of minor elements in sedimentary profiles from Lake Constance [81], Lake Michigan [52-55], Lake Monona/Wisconsin (Lake Minocqua) [48, 49], Lake Washington [72-75]~ and Lake Erie [66]. Data of background and maximum value in ppm. F=factor of increase Lake Constance BackMax ground value Zinc Chromium Nickel Copper Lead Arsenic Mercury Cadmium
124 50 55 30 19 n.d. 0.2 0.21
380 153 50 34 52 0.8 0.68
Lake Michigan
F 3 3 1 1 3 4 3
BackMax ground value 129 77 54 44 40 11 0.04 n.d.
317 85 44 75 145 22 0.2
Lake Monona
F 2.5
1 1 1.5 3.5
2 5
BackMax ground value 15 7 34 22 14 (2 0.24 2.5
from the corresponding inflowing rivers (C, D) are compared in the form of quotients which are listed in increasing order from left to right. The similar increases in the chromium, nickel, cobalt, and iron levels in the western zone (evident from the B:A quotients) indicate that the enrichment is largely due to the lithology of the catchment area (basic intrusive rocks). On the other hand, the enriched mercury, lead, cadmium, and zinc values encountered in the southeastern zone of the Hartbeespoort Dam are most probably due to civilizational effects. The high zinc and lead values can possibly be ascribed to domestic waste effluents whilst the cadmium and, in particular, the mercury enrichment could have resulted from industrial discharges. Similar investigations of metal pollutants in surface lake sediments were reported from highly industrialized areas in Europe (Sweden [21-25], Finland [26], Poland [27], Germany [28, 29], Switzerland [30], Italy [31]), and North America [3243] and have also been carried out in other parts of the world, e.g., in Australia [44] and the Philippines [45]. It should be noted, however, that natural mechanisms may also affect local concentrations of metals in lake sediments [46]. One such example is given by the geothermal mercury "pollution" in rivers and lakes of New Zealand [47].
Metal Contamination Recorded in Dated Sedimentary Cores
As outlined above, geochemical and civilizational anomalies can be overlapped in certain regions, especially in mining areas. Precultural levels of metal contamination in natural systems, therefore, are necessary to evaluate the rate and extent of accumulation of the individual elements due to man's activities. Sediment cores provide a historical record of events occurring in the watershed of a particular lake and enable a reasonable estimation to be made of the background level and the changes in input of an element over an extended period of time; this approach is particularly useful if the rate of sedimentation is known [48]. 466
92 49 50 268 124 51 1.12 4.6
Lake Washington
F 6 7 1.5 12
9 25) 5 2
BackMax ground value 60 u.d. (iron: 15 20 10 0.1 n.d.
230
50 400 200 1.0
Lake Erie BackMax ground value
F 4 1) 3 20 20 10
7 13 40 18 n.d. 0.6 0.04 0.14
42 60 95 58 3.2 0.48 2.4
F 6 4,5 2.5
4 5.5 12 17
The accumulation of minor and trace elements has been analyzed from sediment cores from lakes in Wisconsin [48-51], Lake Michigan [52-58], western Michigan lakes [59, 60], Lake Peoria, Illinois [61], Lake Ontario [62], Lake Erie [63-66], Lake Huron [67, 68], Lake Superior [79], lakes in Quebec and Ontario [70, 71], Lake Washington [72-75], Lake George [76], Lake Windermere in England [77], Swedish lakes [78, 79], Finnish lakes [80], and from lakes in southern Germany [81, 82]. In five of the above-mentioned lakes, namely, Lake Constance, Lake Michigan, Wisconsin lakes, Lake Washington and Lake Erie, the vertical distribution of a large number of elements in core profiles was measured simultaneously. Table 2 summarizes the results of these studies by listing (1) the background levels of minor elements in the deeper part of the cores, (2) the maximum values in the upper layers, and (3) the factors of enrichment as the quotient (2):(1). (It should be noted, however, that these sediments differ widely, both in their composition and texture; the core profile from Lake Monona, for example, exhibits higher portions of carbonate sediments [83] which may be responsible for the lower background contents of some elements.)
Mixed Sewage Inputs: Lake Constance and Lake Michigan Low concentrations of heavy metals are reflected by the concentrations of iron, cobalt, and nickel in most of the lake sediment sequences. These findings are in accordance with the results from highly polluted river sediments, e.g., from the lower Rhine section where the contents of Fe, Co, and Ni are influenced particularly by geochemical factors [84].
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Moderate enhancement of the zinc, lead, mercury, and cadmium values in Lake Constance and Lake Michigan sedimentary cores by factors of 2 to about 7, seems to originate mainly from mixed effluent inputs f r o m industrial, domestic, and agricultural sources. Specification of effluents into domestic and industrial discharges can then be made by the use of additional indicators, e.g., the contents of nitrogen, phosphorus, and organic carbon within the same sedimentary profiles. F6rstner et al. [82] compared the mode of accumulation of N and P with the concentrations of heavy metals in a dated core from central Lake Constance. Similar trends of both zinc/lead and nitrogen/phosphorus concentrations [85] indicate that these substances originate simultaneously in the public sewage system, whereas the increase of cadmium and chromium within the upper layers of the sediment sequences can be explained by industrial emissions, from the electroplating and tanning industries, respectively. Significant increases in the concentrations of zinc, lead, mercury, and cadmium have also been observed in marine sedimentarycores, e.g., in the southern California coastal zone [86], in the Baltic Sea [87], and in the North Sea [88].
Algicides and Herbicides: Wisconsin Lakes Metal pollution resulting from the use of pesticides is demonstrated by examples from lakes in Wisconsin: Copper concentrations in the sediments from Lake M o n o n a increase with depth to a m a x i m u m (434 p p m Cu at 15-20 cm sediment depth) and then decline sharply [48]. These effects are a result of the intermittent treatment of the lakes with copper sulfate to control algal growth during 1918-1944. Since the rates of increase of copper concentrations in sediments from other lakes, e.g., Lake Michigan and Lake Constance, are generally low, local copper accumulations may, as a rule, be taken as indicators of previous administrations of copper for algal control. Arsenic values were not available from Lake M o n o n a and were therefore taken from a study of Lake Minoqua in northeastern Wisconsin [49]. The characteristic increase of arsenic in these sediments was explained by the fact that several Wisconsin lakes had been treated with sodium arsenite to reduce the population of noxious weeds. It is assumed that fertilizers [89] and, above all, household detergents [90] are responsible for the general enrichment of arsenic concentrations, as observed, for example, in the sediments from Lake Michigan [38].
Airborne Particulate Fallout: Lake Washington Characteristic sources of metal pollution are constituted by smelters and by coal-fired power plants [91] and also by heavy city traffic as was first shown by Chow and Patterson [92] in lead-isotope studies on marine sedimentary cores. A n instructive example of atmospheric influences on metal concentrations in lacustrine deposits is given by Crecelius and Piper [73] from sedimentary core studies of Lake Washington: Enhancement of lead Naturwissenschaften 63, 465-470 (1976)
concentrations from a background level of 20 p p m Pb during the 1880s corresponds to the time when land along the western slope of Lake Washington was first developed. The population of Seattle rose from 7,600 in 1880 to over 100,000 in 1900. During this same period, the T a c o m a smelter, 50 km south of Seattle, began its operations. Atmospheric discharges from the smelter are held responsible for a 5- to 10-fold increase in the lead concentrations observed in the sediments up until 1916. Conversion of the smelters from lead to copper production is reflected by a sharp decrease in the lead values in sediments deposited between 1916 and 1920. Above this layer, lead increases strongly to its present values of 400 p p m Pb, probably due to the use of gasoline additives, starting in the 1920s [73]. Similar enrichment of arsenic in present deposits from background concentrations of about 1 0 p p m to more than 200 p p m As is attributed to atmospheric fallout from the T a c o m a smelter [75]. Barnes and Schell [72] estimated that 99% of the current lead input, 93% of the mercury, 68% of the zinc, and 56% of the copper input to Lake Washington sediments originate from aeolian processes. However, since trace metals from atmospheric fallout as well as from fluvial suspended load often occupy inert positions within the particles, their direct effects on both water quality and aquatic life are less detrimental than those arising from solute metal discharges.
Industrial Effluents." Lake Erie Trace and minor element concentrations in sedimentary cores from eight stations in Lake Erie were determined by Walters and coworkers [66]. Table 2 lists the data of a core f r o m the central basin of Lake Erie near Cleveland at the m o u t h of the Cuyahoga river. The example used seems to be representative of a moderate pollution in that area: According to the distribution of mercury in the sedimentary record, early cultural activity might have taken place sometime a r o u n d 1835. M a j o r increases in the values of zinc, arsenic, and copper during 1939-1955 reflect the general growth of industry during World W a r II and the Korean conflict. Strong enrichments of chromium and, in particular, cadmium occurred during the late 1940s, corresponding for the most part to the growth of the Cleveland electroplating industry. It is assumed that the establishment of chemical plants in the Cleveland area in 1949 caused a characteristic increase in the mercury concentrations; the major break in 1955 would then correspond to the opening of the Detrex chlorine-alkali plant near Ashtabula, Ohio [66].
Mercury Poisoning of Lakes The acute toxicity of mercury was evidenced by the catastrophic poisonings in M i n a m a t a and Niigata, Japan, both epidemics originating from mercury-con-
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taminated fish. Meanwhile, the discovery of mercury pollution has been reported in many other parts of the world, e.g., in Scandinavia, Holland, Italy, Canada, United States, Iraq, and Brazil; recently, clinical tests conducted by a Japanese study team in Canadian native reserves indicated various symptoms of mercury poisoning, probably caused by contaminated fish from polluted inland waters [93]. Mercury and many of its compounds exhibit a distinct residueforming tendency in aquatic ecosystems. Ionic mercury is quickly removed from solution and immobilized above its very low background concentrations [94]; microbial processes, however, are capable of converting originally inorganic mercury into more mobile and more toxic methylated forms [65]. Studies carried out by Wood and coworkers [95] and by Jensen and Jernel6v [96] on the factors and mechanisms influencing the biosynthesis of mercury alkyls clearly indicated that the methylation rates are particularly affected by a good supply of nutrients and elevated water temperature. Bioturbation of the upper sediment layers by microorganisms, e.g., tubificids and mussels may prolong the release of methylated mercury from polluted bottom mud for decades [97].
Sources of Mercury Pollution Mercury contamination from sediment and biota has been reported in aquatic systems far removed from characteristic industrial effluents. One likely source of mercury accumulations in remote areas is particulate matter, either from industrial plants burning fossil fuel, or from the smelting of copper, lead, and zinc ores [40]. For example, Lake Saoseo and Lake Tuma in southern Switzerland, both situated at high altitudes, show evidence of atmospheric mercury contamination, which is probably derived from the industrial complexes of nothern Italy [30]. Increases in mercury concentrations often originate from municipal sewer effluents, in particular from plants accepting industrial wastes [98]. Such discharges might be the cause of mercury accumulations in sediments from Lake Trummen in southern Sweden, an extremely eutrophic lake which has been polluted with waste water from the town of V/txj6 [25]. A comprehensive survey of the effects of mercury .pollution in lakes, rivers, and coastal regions was first undertaken in Sweden. After severe cases of birds being poisoned from seed dressings containing methyl mercury dicyandiamide in the early 1960s, the distribution of mercury was investigated in other parts of the environment. These studies led to the detection of industrial effluents as the major source of contamination: 1. Inorganic mercury, chiefly mercury chlorides from the chloralkali industry. 2. Phenylmercury as a preservative for pulp in the cellulose industry, often found in association with fibrous particles. 3. Mercury in various forms, from catalysts in organic chemistry, paint production, manufacture of batteries, from amalgamation of gold and silver etc.
Mercury in Lake Sediments Acute cases of mercury pollution originate mainly from direct industrial emissions. Table 3 lists several examples of mercury pollution in lakes of Sweden 468
Table 3. Maximum concentrations and sources of mercury in lake sediments (examples) Lake/River
Saoseo Trummen V~nern Bj6rken Clay Lake St. Clair River
Max conc. [ppm]
2.2 3.1 10.4 11.2 8.4 60 (2000) Detroit River 86
Source
Ref.
atmospheric fallout municipal waste water chlor~alkali industry wood-pulp mill chlor-alkali industry chlor-alkali industry (Dow Chemical, Sarnia) chlor-alkali industry (Wyandotte Chem., Mich.)
[30] [25] [24] [78] [70] [101] [100] [100]
and North America (other severe cases of mercury pollution have been reported from northern Italy [31] and from Finland [89]; bottom sediments of these lakes containing up to 20 ppm and 170 ppm Hg, respectively): Vfirmlandssj6n, the eastern main basin of Lake Viinern, is one of the most mercury-polluted major lakes in the world [24]. Contamination originated mainly from the chlor-alkali industry of Skoghall. Emissions have successively been reduced from 3,000kg/year during the period 1920-1968, to 500 kg/year during 1969/1971 to approximately 30 kg/ year in 1974 (the latter figure is approximately the minimum loss for mercury cell operations). Lake Bj6rken, situated in central Sweden, is very small (1.8 kin2). The principal source of contamination are wastes from a pulp mill discharging phenyl mercury into the Olofjssj6n River, just upstream from the lake. During the period 1955-1965, about 180 kg Hg was used and probably most of it reached the lake [78]. One of the areas in Canada which is most heavily polluted with mercury is the Wabigoon river system in northwestern Ontario. Fish from Clay Lake, approximately 80 km downstream from Dryden, were found to contain up to 16 ppm of mercury. The source of contamination is believed to be a chlor-a!kali plant at Dryden, which, during the years 1962~59, prior to effluent control, discharged a total of 10,000 kg of mercury in the waste water [70]. The mercury problem in the Great Lakes region roused public interest at the beginning of 1970 when commercial fishing was banned in the waters of the St. Clair River-Lake Erie System. Mercury concentrations of up to 5 ppm, ten times the current U.S. Food and Drug Administration level, were analyzed in fish from Lake St. Clair in April and June 1970 [54]. After an extensive review of all the industries and communities, two main sources of mercury pollution were found in that area [99, 100]: One was Wyandotte Chemical, Michigan, which has a chlorine alkali plant located on the Detroit River, and the other was Dow Chemical Company plant on the St. Clair River at Sarnia, Ontario. In Michigan, chlor-alkali manufacturing, utilizing continuous mercury cathode cells began in 1939 at Wyandotte, where approximately 5 10 kg mercury Naturwissenschaften 63, 465-470 (1976)
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were discharged per day. Chlor-alkali operations at Sarnia commenced in 1947 and discharged an average of 10,000 kg per year before the control system was introduced in 1970.
Restoration of Mercury-Contaminated Lakes Following legislation and enforcement to limit mercury pollution, the mercury concentrations in both sediments and organisms of Lake St. Clair decreased significantly during the period 1970--1974 [102, 103]. There is evidence that mercury-polluted sediments are rapidly eroded from the shallow basin of Lake St. Clair [102]. More serious problems, however, may arise in cases of lower rates of either erosion or sedimentation. With respect to Clay Lake, Armstrong and Hamilton [70] stated that "even with the present control, there is little hope of any rapid improvement in the degree of pollution because of the disturbance of sediments by animals". Several methods for the restoration of mercury-contaminated water bodies have been suggested [25] : (1) dredging of polluted sediments, (2) converting mercury to mercury sulfide with a low methylation rate (anaerobic conditions), (3) binding of Hg to silica or coprecipitation with hydrous Fe and Mn oxides (aerobic conditions), (4) increasing the p H to give volatile dimethyl mercury rather than monomethyl mercury, (5) isolating the polluted sediments from the water body by means of physical barriers, such as polymer film overlays, blankets, or plugs of waste wool, sand, and gravel overlays [104]. As to the latter treatment, which seems to be particularly promising, Thomas [69] reports the case of an unintentional, but very effective covering of mercury-polluted sediments in the Silver Bay of Lake Superior by tailing derived from a taconite processing plant.
Water/Sediment Interactions To conclude the present review of metal pollution in lakes, as determined from sediment studies, it should be noted that potential hazards both for the quality of drinking water and for aquatic life may arise from metal-polluted lake deposits by changes of the water chemistry: Three mechanisms seem~ at present, to be most disadvantageous: Firstly, an increasing input of natural or synthetic chelating substances, by which metal compounds from the sediments are being released into the water [105]. Some of the metal chelates, e.g., copper, nickel, mercury, and cadmium complexed by nitrilotriacetic acid (NTA is used to replace polyphosphates in detergents) have proved highly stable in natural waters [106] and, thus, could endanger the drinking water supply in areas where river and lake water is used for drinking water purposes [107]. Secondly, the changes o f p H values: Acid conditions in lakes originate particularly from mine effluents and from atmospheric SO2 emissions [108]. Low pH values will increase the rates of solubility of most heavy metal compounds and thus decrease the sorption of heavy metals on clay minerals, humic substances, and Naturwissenschaften 63, 465--470 (1976)
hydrous oxides, e.g., of Hg 2§ to hydrous manganese oxide [109] and of Pb to ferric oxide [110]. There are, in addition, pH influences on biosynthetic reactions, for example on the mono-/dimethylmercury equilibrium [25]. Thirdly, the consequences of advanced eutrophication: Co-precipitation and sorption of trace metals on hydrous oxides of Fe, Mn~ and A1 are assumed to represent the principal mechanisms controlling the heavy metal distribution in well-aerated systems [111]. Recent investigations of the mobility of various elements in Lake Washington [72] and Lake Michigan [112] seem to confirm this hypothesis: Mean residence times of A1, Fe, and Mn are two to three orders of magnitude shorter than those of sodium and of the lake water, respectively; heavy metal residence times are also significantly reduced. Oxygen deficiency will mainly affect the redox-sensitive hydrous Fe and Mn oxides, and, indirectly, the cations adsorbed to these substances. Iron, manganese, and other heavy metal compounds [113] are dissolved in the deeper reducing zones of the lake deposits, migrate upward through the sedimentary column, and are subsequently precipitated at the oxidizing sediment/water interface [1 t4]. Further depletion of oxygen may cause dissolution of these hydrous Fe and Mn oxides, thereby releasing toxic metals into the open water. The possible implications of the latter process have been depicted by Edgington and Callender [115] in a study of the minor element geochemistry of Lake Michigan ferromanganese nodules. Since these concretions contain unexpectedly large amounts of arsenic, dissolution would increase the contents of arsenic in the waters of Green Bay, where the nodules are abundant, up to about twice the permissible value for drinking waterl From these examples it becomes evident that the problem of metal pollution in lakes is primarily concerned with complex interactions between particulate matter and water. Further investigations should not only include the separate determination of metal concentrations in both sediments and in solution; but research should be directed at establishing the type of bonding involved between metal species and particulate matter, since the availability of toxic compounds for biological processes depends largely on their characteristic physicochemical behavior. I wish to thank Prof. Dr. German Mfiller for his kind cooperation in our lake investigations, which were financially supported by the Deutsche Forschungsgemeinschaft. 1. Zfillig, H.: Schweiz. Z. Hydrologie 18, 7 (1956). - 2. F6rstner, U., Mfiller, G. : Schwermetalle in Flfissen und Seen als Ausdruck der Umweltverschmutzung. Berlin: Springer 1974. - 3. Allan, R.J.: Geol. Surv. Can. 74-1, 43 (1974). - 4. Allan, R.J.: Abstr. Int. Conf. Heavy Metals in the Environment C-5 (1975). - 5. Coker, W.B., Nichol, I. : Econ. Geol. 70, 202 (1975). - 6. Schmidt, R.C.: Unpubl. Ph.D. thesis, McGill Univ. Montreal (1956), cit. in [5]. - 7. Arnold, R.G. : Saskatchewan Res. Council Geol. Div. Circ 4 (1970). - 8. Dyck, W.: Geol. Surv. Can. 71-1, 70 (1971). - 9. Allan, R.J.: Can. Inst. Min. Met. Bull. 64, 43 (1971). -
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10. Alpin, C.L., Argall, G.O. (eds.): Tailings disposal today. San Francisco : Miller F r e e m a n 1973. - 11. Andrews, R.D. : Proc. Int. Conf. Heavy Metals in the Environment (in press). - 12. F6rstner, U., Wittmann, G.T.W.: G e o f o r u m (in press). - 13. Galbraith, J.H., et al. : G r o u n d Water 10, 33 (1972). - 14. Weatherley, A.H., Dawson, P.: Search 4, 471 (1973). - 15. Wixson, B.G., et al., in: Cycling and Control of M e a l s . EPA, Cincinnati, Febr. 1973, p. 11. - 16. Hutchinson, T.C., et al.: Abstr. 2nd Syrup. Environmental Biogeochemistry (1975). - 17. Qvarfort, U.: Kvart/irgeologiska Avdelningen, Uppsala Univ. 56 (1974). 18. Wittmann, G.T.W., F6rstner, U . : W a t e r SA 1, 76 (1975). 19. Toerien, D.F., et al.: ibid. 1, 15 (1975). - 20. Wittmann, G.T.W., F6rstner, U. :ibid. (in press). - 21. Hasselrot, T.B. : Inst. Freshwater Res., Rept. 48, 102 (1968). - 22. Bengtsson, L., Fleischer, S.: Vatten 1, 73 (1971). - 23. Ahl, T.: Naturvardsverkets Limnol. Unders. Uppsala (1972). - 24. H~kanson, L.: N L U Rept. 80 (1975). - 25. Jernel6v, A., Lann, H.: Environ. Sci. Techn. 7, 712 (1973). - 26. Hinneri, S.: Ann. Univ. T u r k u 56 (1974). - 27. Pasternak, K., Glinski, J. : Acta Hydrobiol. 14, 225 (1972). - 28. Groth, P.: Arch. Hydrobiol. 68, 305 (1971). 29. F6rstner, U., Mfiller, G. : IX. Intern. Sediment. Congr., Nice 2, 57 (1975). - 30. Vernet, J.P., Thomas, R.L.: Eclogae Geol. Helv. 65, 293 (1972); 65, 307 (1972) - 31. Damiani, V., T h o m a s , R.L.: Nature 251, 696 (1974). -- 32. N u s s b a u m , D.G.: Unpubl. P h . D . thesis, Univ. Michigan, A n n A r b o r (1965), cit. in [56]. 33. Landstr6m, I., et al.: NBS Spec. Publ. 312, 353 (1969). 34. D'Itri, F.M., et al.: J. Mar. Techn. Soc. 5, 10 (1971). 35. Copeland, R.A., in: Environmental Mercury Contamination, p. 71 (Hartung]Dinman, eds.). A n n A r b o r Science Publ. 1972 - 36. Konrad, J.G. : ibid., p. 52. - 37. Hesse, J.L., Evans, E.D. : Michigan W a t e r Resources C o m m . (1972). - 38. Seydel, I.S.: Arch. Hydrobiol. 71, 17 (1972). - 39. T h o m a s , R . L : Can. J. Earth Sci. 10, 194 (1973). - 40. Mathis, B.J., Kevern, N.R. : Michigan Inst. Water Res. Rept. 34 (1973). -- 41. Pita, F.W., Hyne, N.J.: Geol. Soc. Amer. Abstr. with Progr. 5, 343 (1973). - 42. Hutchinson, T.C., Fitchko, J.: Proc. Int. Conf. Persistent Chemicals in Aquatic Ecosystems, I. 69 (1974). - 43. Baker-Blocker, A., et al.: Bull. Geol. Soc. Amer. 86, 1358 (1975). - 44. Weatherley, A.H. : Abstr. Int. Conf. Heavy Metals in the Environment. C-308 (1975). - 45. Lesaca, R.M.: ibid. C-105. - 46. F6rstner, U.: N. Jb. Miner. Abh. 118, 268 (1973); Geol. Rdsch. (in press). 47. Weissberg, B.G., Zobel, M.G.: Bull. Environ. Contam. Toxicol. 9, 148 (1973). - 48. Syers, J.K., etal.: Water Air Soil Poll. 2, 105 (1973). - 49. Shukla, S.S., etal.: J. Environ. Qual. 1, 292 (1972). - 50. Bortleson, G.C., Lee, G.F.: Environ. Sci. Techn. 6, 799 (1972). - 51. Iskandar, I.K., Keeney, D.R.: ibid. 8, 165 (1974). - 52. Ruch, R.R., etal.: Illinois Geol. Surv. Environ. Geol. Note 37 (1970). - 53. Shimp, N.F., et al.: ibid. 41 (1971). - 54. Kennedy, E.J., etal.: ibid. 44 (1971). 55. Frye, J.C., Shimp, N.F.: ibid. 56 (1972). - 56. Leland, H.V., et al., in: Trace Metals and Metal-Organic Interactions in Natural Waters, p. 89 (P.C. Singer, ed.). A n n A r b o r Science Publ. 1 9 7 3 . 57. Edgington, D.H., Robbins, J.A. : Abstr. Meet. Amer. Soc. Limnol. Oceanogr. (1974). - 58. Cline, J.T., Chambers, R.L.: ibid. - 59. Dunning, C., et at. : Geol. Soc. Amer. Abstr. with Progr. 7, 751 (1975). - 60. Wheeler, R., Dunning, C.: ibid. 7, 880 (1975). - 61. Collinson, C., Shimp, N.F.: Environ. Geol. Note 56 (1972). - 62. T h o m a s , R.L.: Can. J. Earth Sci. 9, 636 (1972). - 63. Walters, L.J., Jr., et al.: Proc. 15th Conf. Great Lakes Res., p. 306 (1972). - 64. Kovaeik, T.L., Walters, L.J., -
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Jr.: Proc. 16th Conf. Great Lakes Res., p. 252 (1973). - 65. Wolery, T.J., Walters, L.J., Jr.: Proc. 17th Conf. Great Lakes Res., p. 235 (1974). - 66. Walters, L.J., Jr., et al.: ibid. p. 219. - 67. Kemp, A.L.W., et al. : J. Sediment Petrol. 44, 207 (1974). - 68. Kemp, A.L.W., T h o m a s , R.L. : Geol. Soc. Amer. Abstr. with Progr. 7, 795 (19"/5). - 69. Thomas, R.L. : Proc. Int. Conf. Persistent Chemicals in Aquatic Ecosystems I-1 (1974). - 70. A r m strong, F.A.J., Hamilton, A.L., in: Trace Metals and Metal-Organic Interactions in Natural Waters, p. 131 (P.C. Singer, ed.). A n n A r b o r Science Publ. 1973. - 71. Ouellet, M. : Proc. Int. Conf. Heavy Metals in the Environment (in press). - 72. Barnes, R.S., Schell, W.R., in: Cycling and Control of Metals. EPA, Cincinnati, Febr. 1973, p. 45. - 73. Crecelius, E.A., Piper, D.Z. : Environ. Sci. Technol. 7, 1053 (1973). - 74. Schell, W . R . : Abstr. Meet. Amer. Soc. Limnol. Oceanogr. (1974). - 75. Crecelius, E.A. : Limnol. Oceanogr. 20, 441 (1975). - 76. Schoettle, M., Friedman, G.M. : Sedimentol. 21, 478 (1974). - 77. Aston, S.R., et al. : Nature 241,450 (1973). - 78. Axelsson, V., H~kanson, L. : Uppsala Naturgeogr. Inst. Rept. 11 (1971); 14 (1972); 15 (1972); 25 (1973); 35 (1975). - 79. Frederiksson, I., Qvarfort, U. : Geol. F6ren. Stockh. F6rh. 95, 137 (1973). - 80. Alhonen, P., et al. : Vesientutkimuslaitoksen Julkaisurja 7 (1972). - 81. F6rstner, U., Mfiller, G. : Tscherm a k s Miner. Petrol. Mitt. 21, 145 (1974). - 82. F6rstner, U., et al.: Naturwissenschaften 61, 270 (1974). - 83. Sanchez, I., Lee, G . F . : W a t e r Res. 7, 587 (1973). - 84. F6rstner, U., Mfiller, G.: G e o f o r u m 14, 53 (1973). - 85. Wagner, G.: Verh. Int. Vet. Limnol. 18, 475 (1972). - 86. Bruland, K.W., et al.: Environ. Sci. Technol. 8, 425 (1974). - 87. Erlenkeuser, H., et al. : Geochim. Cosmochim. Acta 38, 823 (1974). - 88. F6rstner, U., Reineck, H.-E. : Senckenbergiana marit. 6, 175 (1974). - 89. K a n a m o r i , S., Sugawara, K. : Res. Lab. Nagoya Univ. (1965). -- 90. Angino, E.E., et al. : Science 168, 389 (1970). - 91. Ruch, R.R., et al. : Environ. Geol. N o t e 61 (1973). - 92. Chow, T.J., Patterson, C.C.: Geochim. Cosmochim. Acta 26, 263 (1962). - 93. Ui, J. : Abstr. Int. Conf. Heavy Metals in the Environment B-139 (1975). 94. Gavis, J., Ferguson, J.F.: Water Res. 6, 989 (1972). -- 95. Jensen, S., Jernel6v, A.: Nature 223, 753 (1969). - 96. Wood, J.S., et al.: ibid. 220, 173 (1968). - 97. Jernel6v, A.: Limnol. Oceanogr. 15, 958 (1970). - 98. Evans, P.L., et al. : Water Sewage Works 74 (1973). - 99. Fimreite, N. : Environ. Poll. 1, 119 (1970). - 100. T u m e y , W . G . : J.W.P.C.F. 43, 1427 (1971). - 101. Cline, J.T., et al.: Proc. 16th Conf. Great Lakes Res., p. 223 (1973). - 102. T h o m a s , R.L. : Oral. C o m m u n . Intern. Conf. Heavy Metals Environ. (1975). - 103. Myslik, G.: Oral C o m m u n . ibid. (1975). 104. Bongers, L.H., et al., in: Water Pollution Series, U.S. EPA (1972). - 105. Zitko, V., Carson, W.V.: Chemosphere 3, 113 (1972). - 106. Chau, Y.K., Shiomi, M.T.: Water Air Soil Poll. 1, 149 (1972). - 107. Banat, K., et al.: Chem. Geol. 14, 192 (1974). - 108. Od6n, S., Ahl, T.: Sgrtryck U r Ymer, A r s b o k 1970, p. 103. - 109. Lockwood, R.A., Chen, K.Y. : Environ. Sci. Technol. 7, 1029 (1973). -- 110. Gadde, R.R., Laitinen, H.A.: Environ. Lett. 5, 223 (1973). - 111. Jenne, E.A.: Amer. Chem. Soc. Ser. 73, 337 (1968). - 112. Klein, D.H.: W a t e r Air Soil Poll. 4, 3 (1975). - 113. Cline, J.T., Upchurch, G.B.: Proc. 16th Conf. Great Lakes Res., p. 349 (1973). - 114. Tessenow, U.: Arch. Hydrobiol. Suppl. 47, 325 (1975). - 115. Edgington, D.H., Callender, E. : Earth Planet Sci. Lett. 8, 97 (1970). -
Received January 28, 1976
Naturwissenschaften 63, 4 6 5 - 4 7 0 (1976)
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Heavy metals are one of the most toxic forms of environmental pollutants, constituting a threat both to aquatic life and the quality of drinking water. By analyzing lake sediments, it is possible to determine the provenance, distribution, extent, and also the possible hazards of metal contamination. Sedimentary cores, in particular, provide the means for evaluating the different influences from natural and civilizationaI sources; they represent a historical record of the metal accumulations which have taken place during the past decades as a result of population growth and industrial development.
The significance of sediments in assessing the conditions in aquatic systems [1] is best illustrated by examples f r o m the lacustrine environment. This is a result of the fact that freshwater lakes have formed the center for important cultural developments ever since the earliest days of civilization. As a consequence of increased population and industrialization densities, the threat of pollution has become most acute in such areas [2].
Interferences: Geochemical Background and Man's Impact Environmental as well as explorational geochemistry requires a knowledge of the naturally occurring metals in order to evaluate m a n ' s environmental impact: " B o t h the exploration and environmental geochemist can be looking for the same type of areas, those with high metal concentrations, but obviously from a different m o t i v a t i o n " [3]. This is especially valid for those regions where lake sediment geochemistry is used as a guide to mineralization. A center of recent activities in this respect is Canada, which has a greater area than any other country covered by freshwater lakes [4, 5]. As early as 1956, Schmidt [6] evidenced anomalous metal distribution patterns in lake sediments bordering on areas of mineralization in New Brunswick and Quebec. F r o m studies in Saskatchewan, Arnold [7] and D y c k [8] inferred that at low Naturwissenschaften 63, 465-470 (1976)
sampling density, lake sediments reveal outline metalliferous areas as accurately as stream sediments. Lake sediment geochemistry, at the reconnaissance level, was first described by Allan [9] in a survey covering an area of 3,800 k m 2 in the Coppermine region. Since 1970, thousands of mountain lakes (Appalachia, Cordillera), Arctic lakes, prairie lakes and, in particular, shield lakes have been analyzed both for geochemical exploration and environmental management [4]. The latter aspect of pollution monitoring from lake sediments attracted much interest when mineral exploration was followed by large-scale mining and processing activities in certain areas. As shown by examples from N o r t h America, Australia, and South Africa, these operations affect lake-sediment composition in four main ways [3]: (1) tailings introduced as solids into drainage systems, (2) leachates from onshore tailing piles [10-14], (3) effluents from mines or mining plants [15], and (4) airborne particulates from crushers and smelters [16, 17]. However, the different sources of pollution listed above are generally interrelated, often rendering it impracticable to distinguish between them by means of bulk sediment analysis. The exploration-environmentalgeochemistry"overlap" in lake sediments was exemplarily investigated by Allan [3] in the large nickel-mining area at Sudbury on 26 different examples. Most of the lakes studied had accumulated less than 10 cm of sediment within the last hundred years. Thus, samples taken from below this depth reflect the natural levels which occurred prior to ore exploitation, whereas samples from the 10-cm level indicate the extent of subsequent mining activities. In Kelley Lake, adjacent to the Coppercliff smelters, the nickel content, even at levels down to 15 cm of the sediment, was elevated to such an extent that "it may prove economicallyfeasible to reclaim this or other lakes for their metal content" [3]. An example of the interface of lithologie influences and man-made pollution effects with regard to metal concentrations in lake sediments is described by Wittmann and Ftrstner [18]: The Hartbeespoort Dam, which represents one of the most highly eutrophicated impoundments of South Africa [19], is situated 40 km to the -west of Pretoria between the Daspoort Hills on the southern side and the Magaliesberg mountains constituting the northern boundary. From the western catchment,area, the dam is fed by the Magalies and Skeerpoort rivers, which drain a predominantly agricultural basin. In contrast, the eastern Crocodile river inflow stems from the highly industrialized Kempton Park-Isando area, north of Johannesburg [21]. In Table 1 the heavy-metal contents of characteristic sediment samples from the western (A) and the southeastern (B) zones and
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Table 1. Comparison of metal content in sediments from the western (A) and southeastern (B) zones of the Hartbeespoort Dam and from the corresponding inflow rivers (C, D) Cr Quotient B/A Quotient D/C Enrichment mechanism
Ni
Co
0.56 0.57 0.63 0.52 0.46 0.56 lithologic influences
Fe
Cu
Mn
Hg
Pb
Cd
Zn
0.66 0.63
0.91 0.92
0.97 0.94
1.55 2.48 2.50 4,70 2.25 1.87 man-made pollution
2.55 2.00
Table 2. Distribution of minor elements in sedimentary profiles from Lake Constance [81], Lake Michigan [52-55], Lake Monona/Wisconsin (Lake Minocqua) [48, 49], Lake Washington [72-75]~ and Lake Erie [66]. Data of background and maximum value in ppm. F=factor of increase Lake Constance BackMax ground value Zinc Chromium Nickel Copper Lead Arsenic Mercury Cadmium
124 50 55 30 19 n.d. 0.2 0.21
380 153 50 34 52 0.8 0.68
Lake Michigan
F 3 3 1 1 3 4 3
BackMax ground value 129 77 54 44 40 11 0.04 n.d.
317 85 44 75 145 22 0.2
Lake Monona
F 2.5
1 1 1.5 3.5
2 5
BackMax ground value 15 7 34 22 14 (2 0.24 2.5
from the corresponding inflowing rivers (C, D) are compared in the form of quotients which are listed in increasing order from left to right. The similar increases in the chromium, nickel, cobalt, and iron levels in the western zone (evident from the B:A quotients) indicate that the enrichment is largely due to the lithology of the catchment area (basic intrusive rocks). On the other hand, the enriched mercury, lead, cadmium, and zinc values encountered in the southeastern zone of the Hartbeespoort Dam are most probably due to civilizational effects. The high zinc and lead values can possibly be ascribed to domestic waste effluents whilst the cadmium and, in particular, the mercury enrichment could have resulted from industrial discharges. Similar investigations of metal pollutants in surface lake sediments were reported from highly industrialized areas in Europe (Sweden [21-25], Finland [26], Poland [27], Germany [28, 29], Switzerland [30], Italy [31]), and North America [3243] and have also been carried out in other parts of the world, e.g., in Australia [44] and the Philippines [45]. It should be noted, however, that natural mechanisms may also affect local concentrations of metals in lake sediments [46]. One such example is given by the geothermal mercury "pollution" in rivers and lakes of New Zealand [47].
Metal Contamination Recorded in Dated Sedimentary Cores
As outlined above, geochemical and civilizational anomalies can be overlapped in certain regions, especially in mining areas. Precultural levels of metal contamination in natural systems, therefore, are necessary to evaluate the rate and extent of accumulation of the individual elements due to man's activities. Sediment cores provide a historical record of events occurring in the watershed of a particular lake and enable a reasonable estimation to be made of the background level and the changes in input of an element over an extended period of time; this approach is particularly useful if the rate of sedimentation is known [48]. 466
92 49 50 268 124 51 1.12 4.6
Lake Washington
F 6 7 1.5 12
9 25) 5 2
BackMax ground value 60 u.d. (iron: 15 20 10 0.1 n.d.
230
50 400 200 1.0
Lake Erie BackMax ground value
F 4 1) 3 20 20 10
7 13 40 18 n.d. 0.6 0.04 0.14
42 60 95 58 3.2 0.48 2.4
F 6 4,5 2.5
4 5.5 12 17
The accumulation of minor and trace elements has been analyzed from sediment cores from lakes in Wisconsin [48-51], Lake Michigan [52-58], western Michigan lakes [59, 60], Lake Peoria, Illinois [61], Lake Ontario [62], Lake Erie [63-66], Lake Huron [67, 68], Lake Superior [79], lakes in Quebec and Ontario [70, 71], Lake Washington [72-75], Lake George [76], Lake Windermere in England [77], Swedish lakes [78, 79], Finnish lakes [80], and from lakes in southern Germany [81, 82]. In five of the above-mentioned lakes, namely, Lake Constance, Lake Michigan, Wisconsin lakes, Lake Washington and Lake Erie, the vertical distribution of a large number of elements in core profiles was measured simultaneously. Table 2 summarizes the results of these studies by listing (1) the background levels of minor elements in the deeper part of the cores, (2) the maximum values in the upper layers, and (3) the factors of enrichment as the quotient (2):(1). (It should be noted, however, that these sediments differ widely, both in their composition and texture; the core profile from Lake Monona, for example, exhibits higher portions of carbonate sediments [83] which may be responsible for the lower background contents of some elements.)
Mixed Sewage Inputs: Lake Constance and Lake Michigan Low concentrations of heavy metals are reflected by the concentrations of iron, cobalt, and nickel in most of the lake sediment sequences. These findings are in accordance with the results from highly polluted river sediments, e.g., from the lower Rhine section where the contents of Fe, Co, and Ni are influenced particularly by geochemical factors [84].
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Moderate enhancement of the zinc, lead, mercury, and cadmium values in Lake Constance and Lake Michigan sedimentary cores by factors of 2 to about 7, seems to originate mainly from mixed effluent inputs f r o m industrial, domestic, and agricultural sources. Specification of effluents into domestic and industrial discharges can then be made by the use of additional indicators, e.g., the contents of nitrogen, phosphorus, and organic carbon within the same sedimentary profiles. F6rstner et al. [82] compared the mode of accumulation of N and P with the concentrations of heavy metals in a dated core from central Lake Constance. Similar trends of both zinc/lead and nitrogen/phosphorus concentrations [85] indicate that these substances originate simultaneously in the public sewage system, whereas the increase of cadmium and chromium within the upper layers of the sediment sequences can be explained by industrial emissions, from the electroplating and tanning industries, respectively. Significant increases in the concentrations of zinc, lead, mercury, and cadmium have also been observed in marine sedimentarycores, e.g., in the southern California coastal zone [86], in the Baltic Sea [87], and in the North Sea [88].
Algicides and Herbicides: Wisconsin Lakes Metal pollution resulting from the use of pesticides is demonstrated by examples from lakes in Wisconsin: Copper concentrations in the sediments from Lake M o n o n a increase with depth to a m a x i m u m (434 p p m Cu at 15-20 cm sediment depth) and then decline sharply [48]. These effects are a result of the intermittent treatment of the lakes with copper sulfate to control algal growth during 1918-1944. Since the rates of increase of copper concentrations in sediments from other lakes, e.g., Lake Michigan and Lake Constance, are generally low, local copper accumulations may, as a rule, be taken as indicators of previous administrations of copper for algal control. Arsenic values were not available from Lake M o n o n a and were therefore taken from a study of Lake Minoqua in northeastern Wisconsin [49]. The characteristic increase of arsenic in these sediments was explained by the fact that several Wisconsin lakes had been treated with sodium arsenite to reduce the population of noxious weeds. It is assumed that fertilizers [89] and, above all, household detergents [90] are responsible for the general enrichment of arsenic concentrations, as observed, for example, in the sediments from Lake Michigan [38].
Airborne Particulate Fallout: Lake Washington Characteristic sources of metal pollution are constituted by smelters and by coal-fired power plants [91] and also by heavy city traffic as was first shown by Chow and Patterson [92] in lead-isotope studies on marine sedimentary cores. A n instructive example of atmospheric influences on metal concentrations in lacustrine deposits is given by Crecelius and Piper [73] from sedimentary core studies of Lake Washington: Enhancement of lead Naturwissenschaften 63, 465-470 (1976)
concentrations from a background level of 20 p p m Pb during the 1880s corresponds to the time when land along the western slope of Lake Washington was first developed. The population of Seattle rose from 7,600 in 1880 to over 100,000 in 1900. During this same period, the T a c o m a smelter, 50 km south of Seattle, began its operations. Atmospheric discharges from the smelter are held responsible for a 5- to 10-fold increase in the lead concentrations observed in the sediments up until 1916. Conversion of the smelters from lead to copper production is reflected by a sharp decrease in the lead values in sediments deposited between 1916 and 1920. Above this layer, lead increases strongly to its present values of 400 p p m Pb, probably due to the use of gasoline additives, starting in the 1920s [73]. Similar enrichment of arsenic in present deposits from background concentrations of about 1 0 p p m to more than 200 p p m As is attributed to atmospheric fallout from the T a c o m a smelter [75]. Barnes and Schell [72] estimated that 99% of the current lead input, 93% of the mercury, 68% of the zinc, and 56% of the copper input to Lake Washington sediments originate from aeolian processes. However, since trace metals from atmospheric fallout as well as from fluvial suspended load often occupy inert positions within the particles, their direct effects on both water quality and aquatic life are less detrimental than those arising from solute metal discharges.
Industrial Effluents." Lake Erie Trace and minor element concentrations in sedimentary cores from eight stations in Lake Erie were determined by Walters and coworkers [66]. Table 2 lists the data of a core f r o m the central basin of Lake Erie near Cleveland at the m o u t h of the Cuyahoga river. The example used seems to be representative of a moderate pollution in that area: According to the distribution of mercury in the sedimentary record, early cultural activity might have taken place sometime a r o u n d 1835. M a j o r increases in the values of zinc, arsenic, and copper during 1939-1955 reflect the general growth of industry during World W a r II and the Korean conflict. Strong enrichments of chromium and, in particular, cadmium occurred during the late 1940s, corresponding for the most part to the growth of the Cleveland electroplating industry. It is assumed that the establishment of chemical plants in the Cleveland area in 1949 caused a characteristic increase in the mercury concentrations; the major break in 1955 would then correspond to the opening of the Detrex chlorine-alkali plant near Ashtabula, Ohio [66].
Mercury Poisoning of Lakes The acute toxicity of mercury was evidenced by the catastrophic poisonings in M i n a m a t a and Niigata, Japan, both epidemics originating from mercury-con-
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taminated fish. Meanwhile, the discovery of mercury pollution has been reported in many other parts of the world, e.g., in Scandinavia, Holland, Italy, Canada, United States, Iraq, and Brazil; recently, clinical tests conducted by a Japanese study team in Canadian native reserves indicated various symptoms of mercury poisoning, probably caused by contaminated fish from polluted inland waters [93]. Mercury and many of its compounds exhibit a distinct residueforming tendency in aquatic ecosystems. Ionic mercury is quickly removed from solution and immobilized above its very low background concentrations [94]; microbial processes, however, are capable of converting originally inorganic mercury into more mobile and more toxic methylated forms [65]. Studies carried out by Wood and coworkers [95] and by Jensen and Jernel6v [96] on the factors and mechanisms influencing the biosynthesis of mercury alkyls clearly indicated that the methylation rates are particularly affected by a good supply of nutrients and elevated water temperature. Bioturbation of the upper sediment layers by microorganisms, e.g., tubificids and mussels may prolong the release of methylated mercury from polluted bottom mud for decades [97].
Sources of Mercury Pollution Mercury contamination from sediment and biota has been reported in aquatic systems far removed from characteristic industrial effluents. One likely source of mercury accumulations in remote areas is particulate matter, either from industrial plants burning fossil fuel, or from the smelting of copper, lead, and zinc ores [40]. For example, Lake Saoseo and Lake Tuma in southern Switzerland, both situated at high altitudes, show evidence of atmospheric mercury contamination, which is probably derived from the industrial complexes of nothern Italy [30]. Increases in mercury concentrations often originate from municipal sewer effluents, in particular from plants accepting industrial wastes [98]. Such discharges might be the cause of mercury accumulations in sediments from Lake Trummen in southern Sweden, an extremely eutrophic lake which has been polluted with waste water from the town of V/txj6 [25]. A comprehensive survey of the effects of mercury .pollution in lakes, rivers, and coastal regions was first undertaken in Sweden. After severe cases of birds being poisoned from seed dressings containing methyl mercury dicyandiamide in the early 1960s, the distribution of mercury was investigated in other parts of the environment. These studies led to the detection of industrial effluents as the major source of contamination: 1. Inorganic mercury, chiefly mercury chlorides from the chloralkali industry. 2. Phenylmercury as a preservative for pulp in the cellulose industry, often found in association with fibrous particles. 3. Mercury in various forms, from catalysts in organic chemistry, paint production, manufacture of batteries, from amalgamation of gold and silver etc.
Mercury in Lake Sediments Acute cases of mercury pollution originate mainly from direct industrial emissions. Table 3 lists several examples of mercury pollution in lakes of Sweden 468
Table 3. Maximum concentrations and sources of mercury in lake sediments (examples) Lake/River
Saoseo Trummen V~nern Bj6rken Clay Lake St. Clair River
Max conc. [ppm]
2.2 3.1 10.4 11.2 8.4 60 (2000) Detroit River 86
Source
Ref.
atmospheric fallout municipal waste water chlor~alkali industry wood-pulp mill chlor-alkali industry chlor-alkali industry (Dow Chemical, Sarnia) chlor-alkali industry (Wyandotte Chem., Mich.)
[30] [25] [24] [78] [70] [101] [100] [100]
and North America (other severe cases of mercury pollution have been reported from northern Italy [31] and from Finland [89]; bottom sediments of these lakes containing up to 20 ppm and 170 ppm Hg, respectively): Vfirmlandssj6n, the eastern main basin of Lake Viinern, is one of the most mercury-polluted major lakes in the world [24]. Contamination originated mainly from the chlor-alkali industry of Skoghall. Emissions have successively been reduced from 3,000kg/year during the period 1920-1968, to 500 kg/year during 1969/1971 to approximately 30 kg/ year in 1974 (the latter figure is approximately the minimum loss for mercury cell operations). Lake Bj6rken, situated in central Sweden, is very small (1.8 kin2). The principal source of contamination are wastes from a pulp mill discharging phenyl mercury into the Olofjssj6n River, just upstream from the lake. During the period 1955-1965, about 180 kg Hg was used and probably most of it reached the lake [78]. One of the areas in Canada which is most heavily polluted with mercury is the Wabigoon river system in northwestern Ontario. Fish from Clay Lake, approximately 80 km downstream from Dryden, were found to contain up to 16 ppm of mercury. The source of contamination is believed to be a chlor-a!kali plant at Dryden, which, during the years 1962~59, prior to effluent control, discharged a total of 10,000 kg of mercury in the waste water [70]. The mercury problem in the Great Lakes region roused public interest at the beginning of 1970 when commercial fishing was banned in the waters of the St. Clair River-Lake Erie System. Mercury concentrations of up to 5 ppm, ten times the current U.S. Food and Drug Administration level, were analyzed in fish from Lake St. Clair in April and June 1970 [54]. After an extensive review of all the industries and communities, two main sources of mercury pollution were found in that area [99, 100]: One was Wyandotte Chemical, Michigan, which has a chlorine alkali plant located on the Detroit River, and the other was Dow Chemical Company plant on the St. Clair River at Sarnia, Ontario. In Michigan, chlor-alkali manufacturing, utilizing continuous mercury cathode cells began in 1939 at Wyandotte, where approximately 5 10 kg mercury Naturwissenschaften 63, 465-470 (1976)
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were discharged per day. Chlor-alkali operations at Sarnia commenced in 1947 and discharged an average of 10,000 kg per year before the control system was introduced in 1970.
Restoration of Mercury-Contaminated Lakes Following legislation and enforcement to limit mercury pollution, the mercury concentrations in both sediments and organisms of Lake St. Clair decreased significantly during the period 1970--1974 [102, 103]. There is evidence that mercury-polluted sediments are rapidly eroded from the shallow basin of Lake St. Clair [102]. More serious problems, however, may arise in cases of lower rates of either erosion or sedimentation. With respect to Clay Lake, Armstrong and Hamilton [70] stated that "even with the present control, there is little hope of any rapid improvement in the degree of pollution because of the disturbance of sediments by animals". Several methods for the restoration of mercury-contaminated water bodies have been suggested [25] : (1) dredging of polluted sediments, (2) converting mercury to mercury sulfide with a low methylation rate (anaerobic conditions), (3) binding of Hg to silica or coprecipitation with hydrous Fe and Mn oxides (aerobic conditions), (4) increasing the p H to give volatile dimethyl mercury rather than monomethyl mercury, (5) isolating the polluted sediments from the water body by means of physical barriers, such as polymer film overlays, blankets, or plugs of waste wool, sand, and gravel overlays [104]. As to the latter treatment, which seems to be particularly promising, Thomas [69] reports the case of an unintentional, but very effective covering of mercury-polluted sediments in the Silver Bay of Lake Superior by tailing derived from a taconite processing plant.
Water/Sediment Interactions To conclude the present review of metal pollution in lakes, as determined from sediment studies, it should be noted that potential hazards both for the quality of drinking water and for aquatic life may arise from metal-polluted lake deposits by changes of the water chemistry: Three mechanisms seem~ at present, to be most disadvantageous: Firstly, an increasing input of natural or synthetic chelating substances, by which metal compounds from the sediments are being released into the water [105]. Some of the metal chelates, e.g., copper, nickel, mercury, and cadmium complexed by nitrilotriacetic acid (NTA is used to replace polyphosphates in detergents) have proved highly stable in natural waters [106] and, thus, could endanger the drinking water supply in areas where river and lake water is used for drinking water purposes [107]. Secondly, the changes o f p H values: Acid conditions in lakes originate particularly from mine effluents and from atmospheric SO2 emissions [108]. Low pH values will increase the rates of solubility of most heavy metal compounds and thus decrease the sorption of heavy metals on clay minerals, humic substances, and Naturwissenschaften 63, 465--470 (1976)
hydrous oxides, e.g., of Hg 2§ to hydrous manganese oxide [109] and of Pb to ferric oxide [110]. There are, in addition, pH influences on biosynthetic reactions, for example on the mono-/dimethylmercury equilibrium [25]. Thirdly, the consequences of advanced eutrophication: Co-precipitation and sorption of trace metals on hydrous oxides of Fe, Mn~ and A1 are assumed to represent the principal mechanisms controlling the heavy metal distribution in well-aerated systems [111]. Recent investigations of the mobility of various elements in Lake Washington [72] and Lake Michigan [112] seem to confirm this hypothesis: Mean residence times of A1, Fe, and Mn are two to three orders of magnitude shorter than those of sodium and of the lake water, respectively; heavy metal residence times are also significantly reduced. Oxygen deficiency will mainly affect the redox-sensitive hydrous Fe and Mn oxides, and, indirectly, the cations adsorbed to these substances. Iron, manganese, and other heavy metal compounds [113] are dissolved in the deeper reducing zones of the lake deposits, migrate upward through the sedimentary column, and are subsequently precipitated at the oxidizing sediment/water interface [1 t4]. Further depletion of oxygen may cause dissolution of these hydrous Fe and Mn oxides, thereby releasing toxic metals into the open water. The possible implications of the latter process have been depicted by Edgington and Callender [115] in a study of the minor element geochemistry of Lake Michigan ferromanganese nodules. Since these concretions contain unexpectedly large amounts of arsenic, dissolution would increase the contents of arsenic in the waters of Green Bay, where the nodules are abundant, up to about twice the permissible value for drinking waterl From these examples it becomes evident that the problem of metal pollution in lakes is primarily concerned with complex interactions between particulate matter and water. Further investigations should not only include the separate determination of metal concentrations in both sediments and in solution; but research should be directed at establishing the type of bonding involved between metal species and particulate matter, since the availability of toxic compounds for biological processes depends largely on their characteristic physicochemical behavior. I wish to thank Prof. Dr. German Mfiller for his kind cooperation in our lake investigations, which were financially supported by the Deutsche Forschungsgemeinschaft. 1. Zfillig, H.: Schweiz. Z. Hydrologie 18, 7 (1956). - 2. F6rstner, U., Mfiller, G. : Schwermetalle in Flfissen und Seen als Ausdruck der Umweltverschmutzung. Berlin: Springer 1974. - 3. Allan, R.J.: Geol. Surv. Can. 74-1, 43 (1974). - 4. Allan, R.J.: Abstr. Int. Conf. Heavy Metals in the Environment C-5 (1975). - 5. Coker, W.B., Nichol, I. : Econ. Geol. 70, 202 (1975). - 6. Schmidt, R.C.: Unpubl. Ph.D. thesis, McGill Univ. Montreal (1956), cit. in [5]. - 7. Arnold, R.G. : Saskatchewan Res. Council Geol. Div. Circ 4 (1970). - 8. Dyck, W.: Geol. Surv. Can. 71-1, 70 (1971). - 9. Allan, R.J.: Can. Inst. Min. Met. Bull. 64, 43 (1971). -
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Received January 28, 1976
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