Envlron. Sci. Technol. 1982, 16, 23-30
Spurny, K. R.; Lodge, J. P.; Frank, E. R.; Sheesley, D. C. Environ. Sci. Technol. 1969, 5, 453. Stern, A. C. “Air Pollution, Air Quality Management”, 3rd Ed.; Academic Press: New York, 1977; Vol. V. p 521. Zimdahl, R. L.; Skogerboe, R. K. Environ. Sci. Technol. 1977,11, 1202. Klein, H. D. Environ. Sci. Technol. 1972, 6, 560-1. John. M. K. Environ. Sci. Technol. 1971. 5. 1199-203. Obon, K. W.; Skogerboe, R. K. Environ. Sci. Technol. 1975, 9, 227-31. Instituto -National de Sismologia, Vulcanologia, Meteorologia, e Hidrohgia, Libro de Registros ClimBticos, SecciBn de Climatologia, 1980. Chisolm, J. J. Sci. Am. 1977, 224, 15-20. Dzubay, T. G.; Stevens, R. K. Environ. Sci. Technol. 1975, 9, 663-7. Lee, R. E.; Goransos, S. S.; Enrione, R. E.; Morgan, G. B. Enuiron. Sci. Technol. 1972, 6, 1025-30. Kleinman, M. T.; Pasternack, B. S.; Eisebund, M.; Kneip, T. J. Environ. Sci. Technol. 1980, 14, 64. Rhodes, J. R.; Pradzynski, A. H.; Hunter, C. B. Environ. Sci. Technol. 1972, 6, 922-7. Robles, E. B.S. Dissertation, Universidad de San Carlos de Guatemala, Guatemala, 1977. Lundgren, D. A. In “Aerosols and Atmospheric Chemistry”; Hidy, G. M., Ed.; Academic Press: New York, 1972; p 266. Moriber, G. “Environmental Science”; Allyn and Bacon: Boston, MA, 1974. Turk, A.; Turk, J.; Wittes, J. “Ecologia-ContaminaciBnMedio Ambiente”; Editorial Interamericana, Mgxico, 1973; p 96-101. Chem. Eng. News 1980,27,12. Waldbott, G. K. “Health Effects of Environmental Pollutants”; Mosby: St. Louis, MO, 1973; p 59.
15% of the gasoline used was lead free; in 1980, the unleaded contribution was 50%; and by 1990, it is expected that 80% of the gasoline used will be unleaded (21). The decrease that has occurred in the lead levels in gasoline in the United States has had as a consequence an effective decrease in airborne lead levels (15). Finally, it is interesting to mention the fact that pollutants dangerous to human health have a greater effect on populations whose general physical state is not optimal. For example, a deficiency in vitamins, poor nutritional habits, and malnutrition are factors that increase the sensitivity of a population toward the diseases caused by atmospheric pollution (22). Since the airborne lead levels found in Guatemala City are high and there is a significant section of the population, especially children of low socioeconomic strata, in a poor nutritional state, it is even more urgent in this country to study and control all gaseous (NO,, SOz, COz, etc.) and particulate atmospheric pollutants. In this manner we can contribute to the improvement of the environment and the health of our population. Acknowledgments
We thank one of the referees for bringing to our attention an accurate method of calibrating our flowmeter. Literature Cited (1) Stong, C. L. Sci. Am. 1971,225, 791. (2) “Analytical Methods for Atomic Absorption Spectrometry”; Perkin-Elmer: Norwalk, CT, 1973; p 10-1, Method PC-3. (3) Flocchini, R. G.; Cahill, T. A,; Shadoan, D. J.; Lange, S. J.; Eldred, R. A.; Feeney, P. J.; Wolfe, G. W. Environ. Sci. Technol. 1976, 10, 77. (4) Whitby, K. T.; Husar, R. B.; Liu, B. Y. In “Aerosols and Atmospheric Chemistry”; Hidy, G. M., Ed.; Academic Press: New York, 1972; p 250.
Received for review December 22, 1980. Revised Manuscript Received J u n e 19, 1981. Accepted August 17, 1981.
Anthropogenically Derived Changes in the Sedimentary Flux of Mg, Cr, Ni, Cu, Zn, Hg, Pb, and P in Lough Neagh, Northern Ireland Brian Rlppey,” Robert J. Murphy,+ and Stlrk W. Kyle?
Limnology Laboratory, The New University of Ulster, Traad Point, Drumenagh, Magherafelt, Northern Ireland The concentration-depth behavior of Mg, Cr, Ni, Cu, Zn, Hg, Pb, and P in three sediment cores from a central site in Lough Neagh, Northern Ireland, was examined for changes in the sedimentary flux of these elements. Two main periods of change were found. A change in the catchment erosion-leaching regime in the 17th century, caused by widespread and comprehensive woodland clearance, produced increased sedimentary Mg, Cu, and P b concentrations. A second and larger change occurred after about 1880 A.D. Cr, Cu, Zn, Hg, Pb, P, and, to a lesser extent, Ni concentrations increase toward the sediment surface. Differing P and trace-metal profiles, a comparison of the estimated anthropogenic sedimentary flux with background atmospheric contributions, and a general comparison with other situations all suggest that background atmospheric sources make a substantial contribution to the more recent Cu, Zn, Hg, and P b sedimentary contamination. The trace-metal contamination of Lough Neagh is part of a global pattern. Introduction There have been many reports of increases in the flux + Agricultual and Food Chemistry Research Division, Department of Agriculture, New Forge Lane, Belfast, Northern Ireland
0013-936X/82/0916-0023$01.25/0
of certain trace metals to the sediments of aquatic systems in industrial, in rural, and in remote mountain areas (Table I). Indeed, with increased earth surface fluxes observed in polar regions (19-21), and indications for geochemical models (22-24), there is enough evidence to support the argument that the anthropogenic contribution to the earth surface flux of certain trace metals is important geochemically. We report results for some trace metals in sediment cores from Lough Neagh, Northern Ireland, which indicate that the anthropogenic component of the sedimentary flux of Cr, Ni, Cu, Zn, Hg, Pb, and P is important in this predominantly rural environmental setting. Study Area Lough Neagh is a large, relatively shallow, eutrophic lake with a mainly rural catchment (Table 11). In the catchment there are six towns with over 10000 inhabitants and three with over 20000. Battarbee (25),Smith (26),and the Lough Neagh Working Group (27) give full details of the geographical, agricultural, and urban characteristics of the catchment. The present eutrophic condition of the lake (28) is the result of cultural influences which are believed to have become progressively stronger over the last 100 years or so (25). Phosphorus is the limiting nutrient in Lough Neagh (29),and it was increases of lake
0 1981 American Chemical Society
Environ. Sci. Technoi., Vol. 16,
No. 1, 1982 23
Table I. Sediment Trace-Metal Contamination in Aquatic Systems aquatic system location ocean Baltic Southern California Tokyo and Osako Bays estuary Chesapeake, U.S.A. Narraganset, U. S.A. Saanich, U.S.A. lake Ammersee, West Germany Canodarago, Champlain, and Sylvan, New York, U.S.A. Constance Foundary Cove, New York, U.S.A. Great Lakes Tokyo Palace Moat, Japan Thompson Canyon pond, California, U.S.A. Washington, U.S.A. Whitney and Saltonstall, Connecticut, U.S.A. Windermere, U.K. Wisconsin Lakes, U.S.A. Woodhull, New York, U.S.A.
contaminants Cu, Zn, Cd, Pb V, Cr, Mo, Cu, Ag, Zn, Cd, Pb Cr, Zn, Cu, Pb Cu, Zn, Pb Cr, Cu, Zn, Cd, Pb Cu, Zn, Cd, Hg Cr, Zn, Cd, Pb Cu, Zn, Pb Cr, Zn, Cd, Pb Cu, Zn, Pb Cu, Zn, Cd, Hg, Pb Cr, Co, Ni, Cu, Ag, Zn, Cd, Pb Pb Cu, Zn, Hg, Pb, As, Sb V, Ni, Cu, Ag, Zn, Cd, Pb Cu, Zn, Hg, Pb Cr, Cu, Zn, Cd, Hg, Pb V, Cr, Cu, Ag, Au, Zn, Cd, Pb, Sb
phosphorus loading, first with the introduction of piped sewage disposal and then with the use of phosphate-rich detergents, that caused and still maintains the eutrophic state (25, 26).
Methods One-meter sediment cores were taken with a Mackereth corer (30) from a site in west-central Lough Neagh. These were B41 taken in 1973 and Ba and Bb in 1978. All cores were stored at 4 "C until they were sectioned. They were extruded by using the method outlined by Mackereth (30) and sliced into either 1-or 2-cm sections. The wet density was calculated after weighing the sediment in a vial of known volume, and the percentage dry weight was determined at 105 "C. Loss on ignition was calculated after combustion in a muffle furnace at 550 "C. Above 6%, loss on ignition is proportional to total organic carbon in geological materials, including sediments (31,32). It is used here as an estimate of sediment organic matter content, and its inverse, the mineral content. The dried sediment samples were rendered soluble by the method of Mackereth (33),using Ultrar reagents and
Table 111. Precision and Accuracy of the Analytical Results cu Zn Ni Cr Precision, Standard Deviation in wg g-l core B41 SDa
120
37
8 33
SDa
no. of duplicates spike concn, Mg g-' 7% recovery SDa no. of trials a
24
39
2 38
6 15
17 7
3 7
9 7
3
Accuracy As % Recovery of Spike
10000 105.0 1.4
4.3
2
5
200 102.8
Standard deviation. Environ. Sci. Technol., Vol. 16, No. 1, 1982
100
103.8
1.5 4
1 2
3 4 5
6
7 8 9
10 11 12 13 14 15 16
17 18
double-distilled water throughout. The metal concentrations of the ,solutions were measured by flame atomic absorption spectrophotometry using a Perkin-Elmer 403 for core B41 and an EEL 240 for cores Ba and Bb. Mercury was determined by the cold-vapor technique (34,35)after digestion of the dry sediment (36),and phosphorus by a colorimetric procedure (37). The accuracy and the precision of the analytical procedures were assessed (38). Precision was assessed by the duplicate method (39),and accuracy by using blanks and measurement of the recovery of metals in "spiked" samples. Table I11 gives the analytical precision as the standard deviation. In all cases blanks were less than one precision standard deviation, except in the case of lead, where it was less than two. Percentage recoveries of spikes (Table 111) indicate acceptable accuracy for the analytical procedures. The three cores were dated by biostratigraphic correlation. Battarbee (25,40) has shown a consistent correlation of diatom zonation among 15 cores from Lough Neagh, and this zonation was used here. The Da/Db zone boundary (ca. 1960) was used for Ba and Bb. Da/Db, Db/Dcl (ca. 1915), Dcl/Dc2 (ca. 1880), and Dc2/Dd (interpolated to be ca. 1700) were used for B41. Estimates of the errors produced during data reduction were made by using the method of propagation of random uncorrelated errors (41). Standard deviations of the primary results were either measured or estimated. Linear regressions were calculated by the normal model I method (42). This is considered suitable as the least-squares regression allowing for errors in both variables reduces to the model I regression when the standard deviations of the variables are constant over their ranges (43). This is a
Table 11. Characteristics of Lough Neagh and Its Catchmenta lake area 383 km2 mean depth 8.9 m hydraulic residence time 1.2-1.7 yr catchment area 4453.4 km2 rural population 149 677 urban population 188 921 a Population figures from 1971 census.
no. of duplicates core Bb
ref
20 0 97.8 1.0
4
20, 99.8
3.1 3
Cd
Hg
P
Pb
0.7
0.048
5
110
37
22
39
10
9 7
120 7
200 96.5 9.2 2
1790 99.1 9.4
20 99.0 2.7
4
0.2 103.0
5.8 7
3
-
I
160
,
,
180 YlCW.1
,
,
1LO
!
’
,
150
180
yiW’
I
5
1
1
yicdi.’
80
1
88
6L M l n t l l l l ‘I.
rngvq-1
Figure 1. Concentrations in dry weight and mineral content in core B41.
J
I
I 80
100
VPCug.’
80
W g
110
1
I I10
I10 y m
’
100 1W
150
w1
700
1 15
20, mgh
15
Figure 2. Concentrations in dry weight in cores Ba and Bb.
reasonable assumption given the precision of the results (Table 111) and the concentration ranges (Figures 1and 2).
Results Accumulation Rate. The Da/Db zone boundaries (ca. 1960) for cores Ba and Bb are 7-8 and 12-13 cm, respectively. Assuming a linear accumulation rate, these results give accumulation rates of 0.44 cm a-l for Ba and 0.74 cm a-1 for Bb. The zone boundaries for B41, Da/Db, 6-8 cm, Db/Dcl (ca. 1915), 18-20 cm, Dcl/Dcz (ca. 1880), 24-26 cm, and Dcz/Dd (ca. 1700), 48-50 cm, give linear accumulation rates in the zones of 0.54 cm a-l in Da, 0.27 cm a-1 in Db, 0.17 cm a-l in Dcl, and 0.13 cm a-l in D c ~ The sediment accumulation rate has increased over the last 300 years. No visible change of sediment type was observed in the cores, however. The precision of the calculated accumulation rate was estimated as follows. The position of the zone boundary was taken to have a standard deviation of 1cm, implying a 95% confidence interval 4 cm wide, and the precision of the dating was taken from zloPb and I4C dating techniques used by Battarbee (25). The standard deviation of the 210pbtechnique after 1900 is 4.5 yr, and the standard deviation of the 14C measurement is 90 yr. Given that accumulation rate is calculated from a depth interval divided by a time interval, the standard deviations of the Da zone accumulation rate are 0.18 cm a-I in Ba, 0.29 cm a-l in Bb, and 0.28 cm a-1 in B41. For zone DcZin B41, it is 0.09 cm 8’. From the accumulation rates a date can be interpolated for any depth in the core. The error of this date was estimated by combining the standard deviation of the position of the zone boundary, assumed to be 1 cm, with
the standard deviation of the accumulation rate calculated above. For zone Da the three cores give standard deviations of 6 or 7 yr and for Dcz in B41 91 yr. Patterns of Concentration Change. The concentrations of Mg, Cu, Pb, Zn, Hg, Cr, Ni, Cd, and P and percentage minerals in core B41 (Figure 1)suggest two main patterns of concentration change: firstly, one with constant concentration from the base of the core to 24-26 cm, with increasing concentration above this (Zn, Hg, and Cr show this behavior); secondly, one with constant concentration from the core base up to 52-54 cm, increasing concentration up to 24-26 cm, and finally increasing at a higher rate above this (Cu, Pb, and Ni show this behavior). The Cd and P concentration profiles do not, however, fall into these two categories. The Cd concentration generally decreases from the core base to the surface, while the P profile is similar to the second category above but does not show the constancy of concentration below 52-54 cm. The profiles for Ba and Bb show constant concentration in the upper 20 cm, except for Ni and P b in Bb and P in Ba and Bb (Figure 2). The decrease of Ni and P b concentrations toward the surface in Bb is not, however, large, and so the concentrations of Cu, Pb, Zn, and Ni in the upper 20 cm of Ba and Bb can be taken to be similar to those in the top 10 cm of B41. The P concentration in Ba and Bb increase toward the surface, like B41, but the concentration in the upper 5 cm is higher than in the surface layer of B41. The Mg profile in B41 is discussed below. Erosive Period, 52-54 to 24-26 cm in B41. Mackereth (32) has shown that changes in the major cation concentration of lake sediments reflect changes in the erosionleaching regime in the catchment. Under conditions of high erosion, removal and transport of soil minerals to the lake sediment is sufficiently rapid so that alteration of the mineral constitution is low. Low erosion rates, however, allow sufficient time for in situ weathering and leaching to impoverish the eventual sedimentary material in major cations. The Mg profile (Figure 1)indicates that a small but detectable increase in erosion rate occurs at 52-54 cm and stabilizes at 44-46 cm. The 53-cm level in B41 is estimated to correspond to 1670 (91) A.D. The Mg concentration of the mineral fraction and the mineral content of the sedimentary material would be expected to increase as a result of increased erosion rates (32). Figure 3 shows that the Mg concentration of the mineral fraction does indeed increase above 52-54 cm but that there is no clear increase in the mineral content of the sediment. The Mg results thus do appear to be interpretable in terms of increased erosion rates in the catchment. Further support for this can be found by comparing Figures 1 and 3. Figure 1shows that the Mg concentration per gram dry weight decreases above 24-26 cm, suggesting a change in the erosion regime. There is, however, no evidence for any major land-use changes in the last 100 years (25). As the constitution of the sedimentary material remains unchanged above 24-26 cm at around 18.5 mg of Mg per gram of minerals, Figure 3 indicates that no major land-use changes have occurred recently. It is the increased additions of organic matter to the sediment above 24-26 cm as a result of eutrophication (44)that reduces the percentage minerals and lowers the Mg concentration when expressed per gram dry weight. The sediment chemical evidence is only one piece of evidence indicating higher erosion rates above 52-54 cm. Battarbee (25) shows that evidence from pollen, diatoms, and accumulation rate all support the development of a Environ. Sci. Technol., Vol. 18, No. 1, 1982
25
X
1
I
ygCug-1-0 6.2.18 mgMgg-1 n.27. r 1 ~ 0 . 5 L .P -0 001 x
x
x X
x
x
X
x
x
X
X X
x
X
X X 0
ygPbg-1:-739*751 mgHgg-1 n z 2 1 , r2:0II.P-0001
1
14
I 80
82
81 Miierolr '//.
86
x
period of accelerated mineral inwash. This was interpreted to be the result of a major and comprehensive land-use change, the clearance of woodlands, beginning about 1600 A.D. The increased erosion rate above 52-54 cm in B41 is related to the difference between the two main patterns of concentration change up to 24-26 cm. Figure 4 shows that up to 24-26 cm there is a significant linear relationship between both Cu and P b and Mg ( P < 0.001). No such relationship is found with Cr, Ni, Zn, Cd, and Hg, for these elements exhibit a generally constant concentration up to 24-26 cm (Figure 1). A slight increase in Ni concentration above 52-54 cm (Figure 1)does, however, suggest behavior similar to that of Cu and Pb. The regression of Ni on Mg has a significance less than that conventionally used for acceptance (0.05 < P < 0.10), so that, although suggestive, these results do not reliably show Ni behavior similar to that of Cu and Pb. The increase in erosion rates above 52-54 cm in B41, thus, not only produces increasing Mg concentration but also relates to increasing Cu and Pb concentration. Surface Increases in Concentration 24-26 to 0 cm. Above 24-26 cm, around 1880 (see above), Cu, Pb, Zn, Hg, Cr,and Ni concentrations all increase toward the surface in B41 (Figure 1). This is illustrated more fully in Figure 5, where all of the concentrations are expressed per gram of minerals. The increases in trace-metal concentration above 24-26 cm (Figure 5) contrast with the constant Mg concentration (Figure 3) and suggest that contamination of the sedimentary material is occurring. The linear regressions between the metals above 24-26 cm (Figure 5) indicate common or related sources of the contamination. The Cd concentration in B41 in constant and at its lowest above 24-26 cm, so little contamination is indicated. The Cu, Pb, Ni, and Zn concentrations in the upper 20 cm of Ba and Bb are similar to those in the top 10 cm of B41 (see above) and confirm that contamination of the upper sediment layers is occurring. P, on the other hand, behaves differently to the trace metals in Ba and Bb and somewhat differently in B41. In Ba and Bb the P concentration increases from the base of the core toward the surface (Figure 2), as it does above 24-26 cm (ca. 1880) in B41 (Figure 2). Although most of the trace metals increase above 24-26 cm in B41 also, the relationship of P to Zn is less precise than for the metals. The coefficient of determination for P/Zn (r2= 0.49, n = 9, concentrations per Environ. Sci. Technol., Vol. 16, No. 1, 1982
x
x x
x
Flgure 3. Relationship between Mg concentration per gram of minerals and the mineral content In core B41: (0)54-78 cm, (0)24-54 cm, and ( X ) 0-24 cm.
26
x
88 X
X X
X
x
1L
x
15 Concentrotion, mgMgg-1
e
.
16
I1
Flgure 4. Relationship between Cu and Pb concentrations and the Mg concentration in dry weight in core B41: (0) 54-78 cm, (0)24-54 cm, and (X) 0-24 cm.
gram of minerals) is lower than the smallest of any of the metal/% coefficients (0.80, Figure 5). Thus, the P profiles in the three cores are somewhat different from those of the trace metals. A contemporary P budget is available for the lake (26), and this shows that 60% of the P known to be available for algal growth reaches the lake through sewage disposal works. Present day per capita figures for P are made up of two components in roughly equal amounts, one from excretion and the other from phosphate-rich household detergents (45). As detergents only came into use in the U.K. after 1950, and their use has only gradually increased (45), superimposed on P loading increases due to expanding sewered population in the catchment, there would be a gradual increase in the per capita contribution sometime after 1950. The P profiles in Ba, Bb, and B41 appear then to reflect the increasing P loadings to the lake over the last 100 years or so. Increased lake productivity due to increased P loading on the lake (46) is reflected by rising loss on ignition (Figure 1; 44), and it may be possible that this increases the trace-metal sedimentation efficiency. This, however, is unlikely to be responsible for the trace-metal profiles, since it is unlikely that the increase in loss on ignition of 44% from 24-26 to 0-2 cm in B41 could cause increases in Cu, Zn, and Hg concentration of go%, 111%, and 280%, respectively, in a system such as a lake which is a naturally efficient trap for trace metals (47-49). Overall, although there are differences in detail between the cores, they do reveal increasing contamination of the upper 20 cm of sediment by Cr,Ni, Cu, Zn, Hg, Pb, and P. The contamination patterns of the six metals in these layers appear to be similar and P reflects increased lake loadings originating in the sewerage system.
/ .
I
I
ir;..
.-
0
/
50
b lo
Concentralion ugPbp-1
.. Flgure 5. Relationship between Pb, Cu, Cr, Cu, Hg, and Ni concentrations and Zn concentration per gram of minerals in core 841.
Sedimentary Fluxes. The sedimentary flux is calculated by using the following equation: flux (mg m-2 a-l) = lOabcd where a = accumulation rate (cm a-l), b = fractional dry weight of wet sediment, c = element concentration (pg g-l ). dry weight), and d = wet sediment density (g ~ m - ~ The nature of the results allows only an order of magnitude calculation to be made, and this is for the mean net sedimentary flux for the upper 10 cm (Table IV). The accumulation rates and estimated standard deviations for the Da zone of the cores, 0.44 (0.18) cm a-1 for Ba, 0.74 (0.29) cm a-1 for Bb, and 0.54 (0.28) cm a-l for B41, were calculated above. The density and the fractional dry weight in the upper 10 cm of Ba and Bb are fairly constant, and so the mean and the standard deviation are calculated (Table IV). No density and dry-weight results are available for B41, so these are taken from a core of similar accumulation rate from the same sampling site. This is SMiii in Figure 5 of Battarbee (25),and the results are given in Table IV. An analysis of variance of density and fractional dry weight shows that there are no significant differences in the upper 25 cm of cores Ba, Bb, and SMiii (P < 0.05), so there is probably little bias introduced into the B41 flux calculation. An analysis of variance may be reliably applied to non-Gaussian results provided the asymmetry is not extreme (42),which seem to be the case (see means and standard deviations in Table 111). The trace-metal concentrations in the upper 10 cm of cores Ba and Bb are constant (Figure 2), and most are fairly constant in the upper 10 cm of B41, so means and standard deviations of the metal concentrations are used in the flux calculation. The background flux is calculated from the background concentration of the elements (see below), the accumulation rate in zone Dc2 of B41 (0.13 (0.09) cm a-l) and the density and fractional dry weight at 50 cm in SMiii (25). The results of the flux calculation, together with the standard deviations, are given in Table IV. The anthropogenic flux is the difference between the recent and background flux. These results are discussed below.
Discussion Background Concentrations. The background metal concentrations a t the sampling site were calculated by using material not subjected to cultural influences (50).
The results are shown in Table V as arithmetic means to allow comparison with the other information, the mean for lacustrine deposits (51) and the average shale (52). Table V reveals no departures of Cu, Zn, and P b background concentrations from the geochemical average. Hg is slightly low, Cr and Ni around twice the average, and Cd a good deal above the average. Forstner (51)has shown that Cr and Ni are enriched in basic rocks, and, as a third to a half of the Lough Neagh catchment is underlain by basic olivine-tholeiitic basalts (53, 54), naturally elevated background Cr and Ni concentrations are to be expected. Lyle (54) and Patterson (55) both measured elevated Cr and Ni concentrations in the Irish basalts, and Cr and Ni are high in stream sediments over the whole basaltic region (56). It seems reasonable to conclude that the background concentrations of Cr and Ni are above the geochemical average at the sampling site because of high allogenic inputs from the basalts in the catchment. Erosive Period. The first detectable perturbation of the natural erosion-leaching regime in the Lough Neagh catchment occurs at 52-54 cm in core B41, during the 17th century. This perturbation was a result of widespread deforestation. The evidence for this comes from pollen, diatom, and accumulation rate results (25),and also from rising sedimentary Mg concentrations (Figure 1). There is also support from a slight but detectable change in the composition of the sedimentary mineral material (Figure 3). The alteration of a lake's sedimentary regime, as a result of deforestation or other similar agricultural change in the catchment, has been recorded elsewhere as increased sedimentary K, Mg, and Na concentrations (7,57,58),and by increased accumulation rates (59,57). Decreased sedimentary concentrations of the major cations as a result of the establishment of forest cover after the end of the Pleistocene have also been recorded (58,60,32). I t thus appears that large changes in the erosion-leaching regime of a catchment can be recorded in the major cation concentration of lake sediment. The increased erosion in the Lough Neagh catchment, beginning in the 17th century, not only causes the sedimentary Mg concentration to rise, but also is coincident with increased Cu and Pb concentrations (Figure 1). Both Cu and P b concentrations correlate with those of Mg Environ. Sci. Technol., Vol. 16, No. 1, 1982
27
hhh
P
OOCD
si6
www
w
rlrlrl
rl
n *CDmmmt-
WCDOrlc-c.I rl
u
d Q
a,
aa
rl
3
8 v1
%
0
E 0 rl Li
a a I
rl
e: Q
9 v3
0
E
.-0 Y
c (
a
d 28
Environ. Sci. Technol., Vol. 16, No. 1, 1982
(Figure 41, but Cr, Ni, Zn, Cd, and Hg do not. Michler et al. (7) also found that sedimentary Cu and Na concentrations increased together in a subalpine Bavarian lake, and they believed that this was connected with deforestation. Pb concentrations, however, did not change at that time. Different weathering behavior of the two groups of metals may be responsible for the differing relationships with Mg and increased erosion rate. Pita and Hyne (49) show that Pb and Zn have different weathering behavior. They reported that Zn was selectively weathered from catchment soils and this resulted in higher Zn/Pb ratios in the lake sediments compared to the soils. In such a situation, because the soil Pb/Zn ratio is higher than that of the lake sediment, an increase in soil erosion rate would increase the sedimentary lead concentrations but alter the Zn little. The behavior of P b and Zn in core B41 during the period of increasing erosion appears to follow such a course (Figures 1 and 2). It may also be suggested from the similarity of the profiles that the weathering and transport behaviors of Cu and P b are broadly similar, as are Cr, Ni, Cd, Hg, and Zn (Figures 1-3). Surface Increases in Concentration. There are increases in Cr, Ni, Cu, Zn, Hg, Pb, and P concentration above 24-26 cm (ca. 1880) in core B41 (Figure 1). The results from Ba and Bb (Figure 2) broadly confirm the Ni, Cu, Zn, and P b contamination of the upper sediment layers, but there are differences in detail. This is not unexpected, as variations in accumulation rate of a factor of 2 can occur even over short distances in apparently uniform basins (40,47,61). The variation may be an order of magnitude over the whole of a large lake (47, 62). Varying and increasing accumulation rates as found in Lough Neagh (25,40),together with changing chemical fluxes, may combine to produce different profiles when the concentrations are expressed in terms of dry weight. However, within one core, qualitative differences in profiles can still be used to indicate different input flux patterns. The characteristics of the surface increases in concentration are now examined in an attempt to discover the dominant sources of the contamination. These could be background atmospheric sources, local atmospheric sources, or waste water (11). Comparison of Trace-Metal and P Concentration Profiles. The qualitative differences in the P and trace-metal profiles in Ba and Bb (Figure 2) indicate somewhat unrelated sources of the contamination. It was suggested above that the P profiles in Ba and Bb reflect increased P loadings to the lake as a result of sewered population increase and the introduction of phosphate-rich detergents during the last 100 years or so. Such a relationship between changing lake P loading and sedimentary P concentration has been observed in other situations (63-65). So, if the trace-metal contamination originated in a similar or related way to P, then it would be expected that the sedimentary metal concentrations would be related to local population changes, as is the case with P. Potential ways in which this might happen are if the metal contamination originated from synthetic materials in the catchment (66) and were transported to the lake through the sewerage system or if they originated from local fuel burning and were transported through the sewerage system and via the atmosphere (67, 68). The fact that P concentration increases toward the surface in Ba and Bb, while the trace metals remain constant, suggests that local sources (waste water and atmospheric) do not dominate the trace-metal contamination pattern. Atmospheric sources appear to make a significant contribution to the sediment trace-metal contamination.
Table V. Background Concentrations in Core B41 in pg g-' Dry Weight Cr Ni Zn
a
mean SDa no. of observations mean for lacustrine deposits, < 2 u av shale Standard deviation.
159 5 26 82 90
145 1
21 103 68
123 8 27 123 95
Cd
Hg
cu
Pb
3.7 0.8 27 0.46 0.3
0.24 0.07 28 0.49 0.4
38 4 14 46 45
31 6 15 31 20
Table VI. Comparison of Anthropogenic Sedimentary Fluxes in mg m-* a-' cu Zn Cd Lough Neagh 43-83 123-238 1.8 Woodhull Lake 6 65 Lake Huron 2-14 7-57 0-0.3 Lake Ontario 22-33 109-121 0.9-3.0 Lake Erie 21-66 93-593 0.8-6.0 Lake Windermere 18 169
Comparison of the Sedimentary and Estimated Background Atmospheric Trace-Metal Fluxes. The calculated recent net sedimentary and background fluxes of trace metals in Lough Neagh allow the anthropogenic flux to be estimated (Table IV). This is compared with an estimate of the background atmospheric flux to assess whether atmospheric contamination originating over a larger area than the Lough Neagh catchment is responsible for the sediment trace-metal contamination (18). Cawse (69-71) conducted a 4-year survey of atmospheric trace elements in Britain. He concluded that an aerosol of uniform elemental composition is circulating in the westerly airstreams of the Northern Hemisphere in the middle and upper troposphere, and he estimated this background atmospheric trace-metal flux. In the absence of local information, Cawse's trace-metal atmospheric deposition rates are used here. One of Cawse's sampling sites had concentrations more typical of urban atmospheres, so this is excluded, and one other site was influenced by a Ni smelter, so it is excluded from the Ni calculation. Mean background atmospheric deposition rates are calculated for six sites and 4 years and are shown in Table IV. Table IV shows that the background atmospheric flux could be a significant component of the anthropogenic sedimentary flux of Cu, Zn, Hg, and Pb in Lough Neagh. As atmospheric deposition takes place over the whole of the lake area, and Lough Neagh has considerable areas of the perimeter and southern part underlain by tracemetal-poor sandy deposits (72, 47, 73),the atmospheric contribution at the sampling site could be higher than indicated in Table IV. In either case, the background atmospheric contribution to the Cr and Ni contamination is small. Comparison of Lough Neagh with Other Situations. Trace-metal contamination of the upper sediment layers has been observed in many other lakes (Table I). The estimated anthropogenic flux of trace metals in Lough Neagh (Table IV) is compared with those found in some other situations (Table VI). Woodhull Lake is in a remote area of the Adirondack Mountains, New York State, and the trace-metal contamination there was considered to be due to background atmospheric sources (18). Lakes Huron, Ontario, and Erie form a series with increasing human impact. Kemp and Thomas (11) considered that the lower anthropogenic fluxes found in Lake Huron reflect background atmospheric contributions and that the higher fluxes in this and the other lakes were due to further contributions from
Hg 0.50
0.2-1.4 0.51-2.2 0.49
Pb 54-90 56 8-4 2 63-169 51-232 99
waste water and local atmospheric sources. In the case of southern Lake Huron and Lake Ontario, the background atmospheric inputs were considered to form a not insubstantial portion of the anthropogenic flux. Nriagu et al. (47) compiled a trace-metal budget for Lake Erie and found that, excluding the Detroit River inputs from Lake Huron, sewage discharges and total atmospheric inputs were about equal for Cu, Zn, and Pb. The trace-metal-contamination characteristics of Lough Neagh are intermediate between those of Lakes Ontario and Erie (Table VI). This suggests that both background atmospheric sources and more local sources (waste water and atmospheric) contribute to the contamination in Lough Neagh but that background atmospheric sources are not unimportant. Zn, Pb, and Cu contamination of recent sediments of Lake Windermere has been reported (16), and the present-day anthropogenic fluxes found are included in Table VI. The table shows that the anthropogenic fluxes of Zn, Pb, Cu, and Hg are similar in Lough Neagh and Windermere, a lake also with an essentially rural catchment. Hamilton-Taylor (16) also suggested that the atmosphere is a major potential source of the elevated trace-metal flux to Lake Windermere sediments. The evidence presented above on which to base a decision about the origins of the measured trace-metal contamination in Lough Neagh is indirect. There may also be problems of interpretation if postdepositional effects, especially for P, have changed (74). However, the evidence of differing P and trace-metal profiles in cores Ba and Bb, comparison of the estimated anthropogenic sedimentary flux with background atmospheric contributions, and comparison with other situations all suggest that background atmospheric sources make a substantial contribution to the trace-metal contamination of Cu, Zn, Hg, and Pb, with contributions also from more local sources (atmosphere and waste water). Local sources, probably waste water, appear to have enhanced the somewhat higher natural fluxes of Cr and Ni in Lough Neagh. In conclusion, it also appears that the trace-metal contamination of Lough Neagh sediments, and the importance of background atmospheric sources, are part of a global pattern (Table I and above). Acknowledgments
We thank R. Battarbee and R. Flower for providing the core material and carrying out the diatom analysis on core B41 and cores Ba and Bb, respectively. We also thank R. Environ. Sci. Technol., Vol. 16, No. 1, 1982
29
B. Wood and R. Battarbee for comments on the manuscript, and G. Mohan and H. McGrogan for help with the analyses. Literature Cited (1) Erlenkeuser, H.; Suess, E.; Willkemm, H. Geochim. Cosmochim. Acta. 1974, 38, 823-42. (2) Bruland, K. W.; Bertine, K. K.; Koide, M.; Goldberg, E. D. Environ. Sci. Technol. 1974, 8 , 425-32. (3) Matsumoto, E.; Yokota, S. Kaguya 1976, 46, 182-4. (4) Goldberg, E. D.; Hodge, V.; Koide, M.; Griffin, J.; Gamble, E.; Bricker, 0. P.; Matisoff, G.; Holdren, G. R., Jr., Baun, R. Geochim. Cosmochim. Acta 1978,42, 1413-26. (5) Goldberg, E. D.; Gamble, E.; Griffin, J. J.; Koide, M. Estuarine Coastal Mar. Sci. 1977,5, 549-61. (6) Matsumoto, E.; Wong, C. S. J. Geophys. Res. 1977, 82, 5477-81. (7) Michler, G.; Simon, K.; Whilhelm, F.; Steinberg, C. Arch. Hydrobiol. 1980, 88, 24-44. (8) Wahlen, M.; Thompson, R. C. Geochem. Cosmochim.Acta 1980, 44, 333-40. (9) Forstner, U.; Muller, G. Tschermaks Mineral. Petrogr. Mitt. 1974,21, 145-63. (19) Bower, P. M.; Simpson, H. J.; Williams, S. C.; Li, Y.-H. Environ. Sei. Technol. 1978, 12, 683-7. (11) Kemp, A. L. W.; Thomas, R. L. Water, Air, Soil Pollut. 1976,-5, 469-90. (12) Goldberg, E. D.; Hodge, V.; Koide, M.; Griffin, J. J. Geochem. J.-1976,10,165-74. (13) Shirahata, H.; Elias, R. W.; Patterson, C. C.; Koide, M. Geochim. Cosmochim. Acta 1980, 44, 149-62. (14) Barnes, R. S.; Schell, W. R. In “Cycling and Control of Metals”; proceedings of an environmental resources conference, National Environmental Research Center, US. EPA, Cincinnati, OH, 1973; pp 45-53. (15) Bertine, K. K.; Mendeck, M. F. Enuiron. Sci. Technol. 1978, 12,201-7. (16) Hamilton-Taylor, J. Environ. Sei. Technol. 1979,13,693-7. (17) Iskandar, I.; Keeney, D. R. Environ. Sci. Techol. 1974,8, 165-70. (18) Galloway, J. N.; Likens, G. E. Limnol. Oceanogr. 1979,24, 427-33. (19) Boutron, C.; Delmas, R. Ambio 1980, 9, 210-5. (20) Herron, M. M.; Langwar, C. C., Jr.; Weiss, H. W.; Cragin, J. H. Geochim. Cosmochim. Acta 1977,41,915-20. (21) Murozumi, M.; Chow, T. J.; Patterson, C. Geochim. Cosmochim. Acta 1969, 33, 1247-91. (22) Lantzy, R. J.; Mackenzie, F. T. Geochim. Cosmochim. Acta 1979,43, 511-26. (23) Mackenzie, F. T.; Wollast, R. In “The Sea”; Goldberg, E. D., Ed.; Wiley-Interscience: New York, 1977; Vol. 6, pp 739-88. (24) Stumm, W., Ed. “Global Chemical Cycles and Their Alterations by Man”; Dahlem Konferenzen: West Berlin, 1977. (25) Battarbee, R. W. Philos. Trans. R. Soc. London,Ser. B 1978, 281, 303-45. (26) Smith, R. V. Water Res. 1977,11, 453-9. (27) Lough Neagh Working Group, Advisory Report; H.M.S.O.: Belfast, 1971; Vol. 2. (28) Wood, R. B.; Gibson, C. E. Water Res. 1973, 7, 173-87. (29) Gibson, C. E.; Wood, R. B.; Dickson, E. L.; Jewson, D. H. Mitt. Znt. Ver. Limnol. 1971, 19, 146-60. (30) Mackereth, F. J. H. Limnol. Oceanogr. 1969,14,145-51. (31) Dean, W. E., Jr. J. Sediment. Petrol. 1974,44, 242-8. (32) Mackereth, F. J. H. Philos. Trans. R. SOC.London, Ser. B 1966,250, 165-213. (33) Mackereth, F. J. H. “Some Methods of Water Analysis for Limnologists”; Freshwater Biological Association Scientific Publication No. 21, 1963. (34) Manning, D. C. At. Absorpt. Newsl. 1970, 9, 109. (35) Hatch, W. R.; Ott, W. L. Anal. Chem. 1968, 40, 2085-7.
30
Environ. Sci. Technol., Vol. 16, No. 1, 1982
Skare, I. Analyst (London) 1972, 97, 148-55. Golterman, H. L.; Clymo, R. S. “Methods for Chemical Analysis of Freshwaters”; Blackwell: Oxford, 1969. Bately, G. E.; Gardner, D. Water Res. 1977, 11, 745-6. Youden, W. J. In “Treatise on Analytical Chemistry”; Kolthoff, J. M., Elving, P. J., Eds.; Wiley: New York, 1959; Vol. 1, pp 47-66. Battarbee, R. W. Verh. Znt. Ver. Limnol. 1978,20,624-9. Bevington, P. R. “Data Reduction and Error Analysis for the Physical Sciences”; McGraw-Hill: New York, 1969. Sokall, R. R.; ROW,F. J. “Biometry”; W. H. Freeman: San Francisco, 1969. Barker, D. R.; L. M. Diana Am. J. Phys. 1974,42,224-7. Hargrave, B. T. J. Fish. Res. Board Can. 1973,30,1317-26. Devey, D. G.; Harkness, N. Water Res. 1973, 7, 35-54. Schindler, D. W. Science 1977, 195, 260-2. Nriagu, J. 0.;Kemp, A. L. W.; Wong, H. K. T.; Harper, N. Geochim. Cosmochim. Acta 1979,43, 247-62. Turekian, K. K. Geochim. Cosmochim. Acta. 1977, 41, 1139-44. Pita, E. W.; Hyne, N. J . Water Res. 1975, 9, 701-6. Hakanson, L. “Sediments As Indicators of Contamination-Investigations in the Four Largest Swedish Lakes”; SNV PM 839/Natarvardsverkets, Limnologiska Undersokning Rapport 92, 1977. Forstner, U. Arch. Hydrobiol. 1977, 80, 172-91. Turekian, K. K.; Wedepohl, K. H. Bull. Geol. SOC.Am. 1961, 72, 175-92. Wilson, H. E. “Regional Geology of Northern Ireland”; H.M.S.O.: Belfast, 1972. Lyle, P. J. Earth Sci. (Dublin) 1980, 2, 137-52. Patterson, E. M. Geochim. Cosmochim. Acta 1951, 2, 283-99. Applied Geochemistry Research Group, “Provisional Geochemical Atlas of Northern Ireland”, Imperial College of Science and Technology Technical Communication No. 60, 1973. Pennington, W. Pol. Arch. Hydrobiol. 1978, 25, 429-37. Adams, R. W.; Duthie, H. C. Int. Rev. GesamtenHydrobiol. 1976, 61, 21-36. Birch, P. B.; Barnes, R. S.; Spyridakis, D. E. Limnol. Oceanogr. 1980, 25, 240-7. Likens, G. E.; Davis, M. B. Verh. Znt. Ver. Limnol. 1975, 19,982-93. Pennington, W.; Cambray, R. S.;Eakins, J. D.; Harkness, D. D. Freshwater Biol. 1976, 6, 317-31. Edgington, D. N.; Robbins, J. A. Environ. Sei. Technol. 1976,10, 266-74. Moss, B. Freshwater Biol. 1980, 10, 261-80. Williams, J. D. H.; Murphy, T. P.; Mayer, T. J . Fish. Res. Board Can. 1976,33, 43019. Edmondson, W. T. Proc. Natl. Acad. Sci. U.S.A. 1974,71, 5093-5. Norton, S. A.; Dubiol, R. F.; Sasseville, D. R.; Davis, R. B. Verh. Int. Ver. Limnol. 1978, 20, 538-45. Bertine, K. K.; Goldberg, E. D. Science 1971,173, 233-5. Lazrus, A.; Lorange, E.; Lodge, J. P., Jr. Environ. Sci. Technol. 1970, 4, 55-8. Cawse, P. A. “A Survey of Atmospheric Trace Elements in the U.K.: Results for 1975”; H.M.S.O.: London, 1976. Cawse, P. A. “A Survey of Atmospheric Trace Elements in the U.K.: Results for 1974”; H.M.S.O.: London, 1975. Cawse, P. A. “A Survey of Atmospheric Trace Elements in the U.k. (1972-73)”; H.M.S.O.: London, 1974. Battarbee, R. W. D.Phi1. Thesis, New University of Ulster, Ulster, Northern Ireland, 1973. Oliver, B. G. Environ. Sei. Technol. 1973, 7, 135-7. Jones, B. F.; Bowser, C. J. In “Lakes: Chemistry, Geology, Physics”; Lerman, A. Ed.; Springer-Verlag: Heidelberg, 1978; pp 179-236.
Received for review February 19,1981. Accepted August 27,1981.
