Paleolimnological evidence of recent acidification in two sudbury (Canada) lakes

The Science of the Total Environment, 67 (1987)53-67 Elsevier Science Publishers B.V., Amsterdam -- Printed in The Netherlands

53

PALEOLIMNOLOGICAL EVIDENCE OF RECENT ACIDIFICATION IN TWO SUDBURY (CANADA) LAKES

SUSHIL S. DIXIT*, ARUNA S. DIXIT* and R. DOUGLAS EVANS Trent Aquatic Research Centre, Trent University, Peterborough, Ontario K9J 7B8 (Canada) (Received February 17th, 1987; accepted May 25th, 1987)

ABSTRACT As a resultof an increasein SO2 emissionsfrom the metal mining and smelting activitiesin the vicinity of Sudbury, severe detrimental effectshave been reported on aquatic ecosystems. However, documentation of the time and rates of lake acidificationare not available.The purpose of the present study is to provide thisinformation by utilizingthe stratigraphicanalysisof sedimentary diatoms from Hannah and Clearwater lakes. The study indicates that, in Hannah Lake, acidificationoccurred soon afterthe roasting of ore startedat Copper Cliffin the 1880s.Between about 1880 and 1975,the inferredlake water pH declined from a high of 6.0 to a low of 4.6.After the lake was neutralizedin 1975 itsmeasured pH increasedfrom 4.3to ~ 7.0.This increasewas also indicatedby an increasein diatom-inferredpH. In Clearwater Lake, acidificationcommenced after 1930. Between about 1930 and 1970 the pH declined from ~ 6.0 to ~ 4.2.Due to reduction in SO2 emissions, no further pH decline has occurred since about 1970. The study indicates that, in addition to the existingbufferingcapacity of the lakes,the lake'sdistance from the point source and past changes in smelting practicesmay have greatly influencedthe onset of acidificationin Sudbury lakes. INTRODUCTION T h e emissions from the metallurgical installations in the Sudbury basin characterize it as an area of major environmental interest. In addition to being recognized as a large point source of SO2 emission, the deposition of Cu, Ni and Fe is significantly higher than in any other industrial environment ( M O E , 1982). In Sudbury, metal related activity began in 1883 w h e n Ni ore w a s first discovered (Howard-White, 1963). O p e n pit roasting started at Copper Cliff in 1888 and the production increased sharply after the opening of Falconbridge Smelter and the installation of a 155 m tall stack at Copper Cliff in the 1920s (Nriagu et al., 1982). T h e very acidic nature of lakes in the Sudbury area w a s observed as early as the late 1950s ( G o r h a m and Gordon, 1960) and by 1971 it w a s s h o w n that lakes located up to a distance of 40 k m from Sudbury had very low p H and high trace metal concentrations (Ontario W a t e r Resources Commission, 1971). T h e emissions from S u d b u r y have resulted in the widespread destruction of forest up to a distance of 5 k m from the smelters ( G o r h a m and Gordon, 1960; Hutchin* Present address: Department of Biology, Queen's University, Kingston, Ontario K7L 3N6, Canada. 0048-9697/87/$03.50

© 1987 Elsevier Science Publishers B.V.

54

son and Whitby, 1977). Severe detrimental effects to other biota are also reported, especially to fish (Keller et al., 1980), phytoplankton (Yan, 1979), zooplankton (Yan and Strus, 1980; Keller and Pitblado, 1984) and aquatic macrophytes (Yan et al., 1985). Although in the vicinity of Sudbury, metal related anthropogenic activity started in the later part of the last century, the exact time of the onset of acidification in lakes is not known because long-term pH records are not available. To reconstruct past lake water pH, the paleolimnological approach using the pH-indicator potential of diatoms has been successfullyused (Davis et al., 1983; Flower and Battarbee, 1983; Charles, 1984). In addition to having a specific pH preference, the microfossils deposited in lake sediments represent a time averaged history of the lake environment (incorporating seasonal variations). The objective of our study is to apply this approach to an investigation of when and to what extent lakes have acidified in the Sudbury region as a result of anthropogenic activity. STUDY AREA AND STUDYLAKES A large area of the Sudbury basin is characterised by a geological environment that is highly resistant to chemical weathering, thus making it vulnerable to inputs of strong acids. Chan et al. (1982) reported that the average pH of precipitation in the Sudbury area is 4.3. Hannah and Clearwater lakes, situated southwest of Sudbury, were selected for the stratigraphic analysis of diatoms (Fig. 1). The lakes are located on the granitic and siliceous rocks of the Canadian Shield. The prevailing wind patterns in the study area are northeasterly in winter and southwesterly in summer (Chan et al., 1982; MOE, 1982). Hannah Lake is polymictic, staying thermally homogenous at all times of the year, whereas Clearwater Lake is dimictic. Selected physical and chemical characteristics of these two lakes are presented in Table 1. MATERIALSAND METHODS In July 1984, sediment cores ( ~ 20-25 cm long) were collected from Hannah and Clearwater lakes using a Kajak-Brinkhurst (K-B) gravity corer. The corer

t

/

is oo

cli~S,~U D~B U R Y ~

Copper

,.,

iCleo rwoter L.

/ ,.00. i

69.,.

\

Fig. I. Location of Hannah and Clearwater lakes.

-

55 TABLE

1

Selected physical and chemical characteristics of Hannah and Clearwater lakes (Dixit, 1986) Characteristics Surface area (ha)

Hannah Lake

Clearwater Lake

23.7

76.5

Watershed area (ha) Maximum depth (m) Elevation (m) Secchi (m) pH Alkalinity ~eq l- i) Conductivity ~uScm2) Total P (pg1-1) Sulfate (ragl- 1)

76 8 306 4.5 6.81 310 308.4 7.9 18.6

340 21 302 10 4.49 - 41 76.4 5 20.6

Cu ~ugl -I) Ni (ugl -~)

22 199

159 195

Fe ~gl -~) A1 Ozg1-1)

53 24

47 310

was modified to hold plexiglass coring tubes of 6.3 cm internal diameter, a size which has been shown to be reliable in obtaining undisturbed cores of soft sediments and showing no compaction to a depth of 20 cm (Evans and Lasenby, 1984). On shore, cores were extruded in a vertical position and sliced. Diatoms were cleaned from the sediment samples using an acid digestion technique (Dixit, 1986) and cleaned subsamples were poured into Battarbee plates (Battarbee, 1973). After all the water had evaporated, the microfossil coated cover slips were mounted on glass slides using Hyrax mounting medium (refractive index 1.63). For each sample a minimum of 600 diatom valves were counted and identified in random fields under oil immersion at 1000 x magnification using a Carl Zeiss microscope equipped with phase contrast. Diatom identifications were based on Hustedt (1931-1959), Huber-Pestalozzi (1942), Cleve-Euler (1951-1955), Nygaard (1956), Patrick and Reimer (1966, 1975), Foged (1981), Germain (1981), van Dam et al. (1981), and the PIRLA Diatom Iconograph (unpublished). The diatom taxa of all the samples counted were assigned a pH indicator value on the basis of published literature (Hustedt, 1939; Nygaard, 1956; Meril~iinen, 1967; Patrick and Reimer, 1966, 1975; Lowe, 1974; Foged, 1981; van Dam et al., 1981; Charles, 1985; Anderson et al., 1986; Dixit and Dickman, 1986; and others) along with their pH-related distribution in the surface sediments (0-0.25 cm) of the 30 Sudbury lakes (Dixit, 1986). The grouping was primarily in accordance with Hustedt's (1939) categories with minor modifications. Since Hustedt assigned the "indifferent" category to diatoms which are equally common on both sides of pH 7, the term indifferent was substituted by circumneutral. The indifferent category was used for only those diatoms which appear to have no specific pH preference. Only a very small number of taxa ( < 5% of the total number of taxa identified) could not be categorized into any specific pH indicator category because of the poor understanding of their

56

TABLE 2 Inferred pH regression equations derived from the surface sediment diatom assemblages of 30 Sudbury lakes (Dixit, 1986) Equation

r2

Standard error

p

1. Index ~ pH = 6.48 - 0.99 log index • 2. Index B pH = 6.75 - 1.18 log index B 3. Multiple reg. pH = 7.44 - 1.99 log ACB - 0.019 ACP - 0.011 CIR + 0.242 log ALP

0.88 0.88

_+0.42 + 0.42

< 0.0001 < 0.0001

0.93

_+0.32

< 0.0001

autecology. Since the p e r c e n t a b u n d a n c e of these individual t a x a was always < 1%, it is h i g h l y u n l i k e l y t h a t t h e y m a y influence the diatom-inferred pH. To provide a q u a n t i t a t i v e i n t e r p r e t a t i o n of the pH h i s t o r y of H a n n a h and C l e a r w a t e r lakes, the i n f e r r e d pH profiles were c o n s t r u c t e d using the multiple r e g r e s s i o n e q u a t i o n (Dixit, 1986) based o n d a t a from 30 S u d b u r y a r e a lakes (Table 2). T h e multiple r e g r e s s i o n t e c h n i q u e was used b e c a u s e of its s u p e r i o r i t y o v e r index a l p h a and index B. S e d i m e n t cores were also a n a l y z e d for Cu and Ni. S e d i m e n t subsamples from the v a r i o u s levels of the cores were dried for 24 h at 90°C and t h e n digested o v e r n i g h t at 70°C using a 2:1 m i x t u r e of H C l a n d HNO8 (Bronson, 1975). Copper and Ni c o n c e n t r a t i o n s were d e t e r m i n e d at the a p p r o p r i a t e w a v e l e n g t h s using a V a r i a n flame a t o m i c a b s o r p t i o n s p e c t r o p h o t o m e t e r . T h e cores w e r e d a t e d using t h e 21°pb method. T h e 21°pb a c t i v i t y in sediment samples was m e a s u r e d following the m e t h o d s of E a k i n s and M o r r i s o n (1978) and E v a n s (1980). T h e dates were c o m p u t e d using the CRS model ( C o n s t a n t R a t e of Supply) of Appleby and Oldfield (1978), with associated assumptions. RESULTS AND DISCUSSION

Hannah Lake B e c a u s e the d i s t r i b u t i o n of algal microfossils is n o t u n i f o r m o v e r the whole lake b o t t o m (Dixit and Evans, 1986), difficulty m a y arise w h e n pH r e c o n s t r u c tions are based on a single sediment core. As discussed by Dixit and E v a n s (1986), due to the polymictic n a t u r e of H a n n a h Lake, the pH profiles of the top 3 cm sections of cores collected from t h e lake b o t t o m w h e r e u n d i s t u r b e d sedim e n t a c c u m u l a t e s yielded similar trends. M o r e o v e r , t h e multiple r e g r e s s i o n pH profiles of these 3 cm cores r e p r o d u c e d the lake's k n o w n pH r e c o r d with reasonable a c c u r a c y . Based on these observations, the diatoms of a sediment core from the a r e a of m a x i m u m d e p t h (8 m) were a n a l y z e d to provide e s t i m a t i o n s of past pH in H a n n a h Lake. Since t h e m e t a l m i n i n g and smelting a c t i v i t y in the S u d b u r y a r e a is well d o c u m e n t e d and r e g u l a r pH m e a s u r e m e n t s h a v e been made since 1973 (Table 3), the reliability of the i n f e r r e d pH h i s t o r y of H a n n a h L a k e c a n be e x a m i n e d further.

57 TABLE3 Past pH measurements of Hannah and Clearwater lakes. 1969-1970 data (Ontario Water Resources Commission, 1971); 1973-1979 data (Yah and Lafrance, 1984); 1980-1983 data (Dillon et al., 1986; MOE, unpublished); 1984 (Dixit, 1986) Year

Hannah Lake

Clearwater Lake

1969-70 1973 1974 1975 1976 1977 1978 1979 1980 1981 1982 1983 1984

4.30 4.31 4.29 7.02 6.67 6.59 6.96 7.07 6.57 6.61 6.85 6.92 6.81

4.15 4.29 4.24 4.33 4.23 4.10 4.40 4.41 4.46 4.43 4.48 4.57 4.49

The unsupported 2Z°pb activity in the core suggests a log-linear relationship with depth for the sediment accumulated since 1922 (Fig. 2). The absence of a mixing zone in the 21°pb profile indicates that the sediment is undisturbed. At the lower end of the core, deviation from log-linearity suggests that there has been changes in the sediment accumulation rates in the past. Therefore, the CRS model (Appleby and Oldfield, 1978) was used to compute the dates. The recent increase in sedimentation rate is most likely related to greater inputs of allochthonous matter from the drainage basin as a result of destruction of the surrounding vegetation by SO2 fumigation (Gorham and Gordon, 1960). Although a total of 102 diatom taxa were identified in the core (Dixit, 1986), 0.00

- 1984 -1981 -1972

-0.50

-1960

o o

-1941 -,s ~3

-1922

~"

U

3E - - 1 . 0 0 -1877

- 1.50

5 10 L e a d - 2 1 0 Activity (dpm/g

50

dry mass)

Fig. 2. Unsupported 2Z°pb activity in core from Hannah Lake. MCDM = midpoint cumulative dry mass of sediment per unit area.

58 I -$

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hl,

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. .o

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.....

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59 only 11 taxa were common (Fig. 3). Before about 1880 there were few floristic changes, but since then major shifts have occurred in Hannah Lake. The first major change in diatom flora occurred between about 1880 and 1975 (Fig. 3). The percent abundance of the common acidobiontic taxon Eunotia exigua greatly increased but then stayed at a near constant level. Significant but gradual increases were also seen in the majority of the common acidophilous taxa (e.g. Frustulia rhomboides, Navicula rnediocris, and Eunotia species). Among the common circumneutral taxa, the counts of Achnanthes minutissirna increased, whereas the counts of Cyclotella cornta and Stauroneis anceps declined. The percent abundance of C. stelligera, a pH-indifferent form, also declined. Since the data presented here represent relative abundances, the decline in abundance of this pH indifferent taxon is most likely due to increases in the relative abundance of acidic diatoms. Anornoeoneis vitrea and Pinnularia subcapitata remained relatively unchanged and Tabellaria flocculosa declined only after 1960 (Fig. 3). The second major diatom shift occurred between about 1975 and 1984 (Fig. 3). It was marked by the decline in the percent abundance of F. rhornboides, N. rnediocris, and acidophilous Eunotia spp., and increases in Anornoeoneis vitrea and Cyclotella stelligera. These changes in the flora appear to be closely associated with the neutralization of the lake in 1975 by the additions of Ca(OH)2 and CaCO8 (MOE, 1982). This provides evidence that, as a result of a major pH increase, a quick response c a n occur in diatoms. As a result of lime treatment in Lake Valkelampi, Simola (1986) also observed an immediate change in diatom species composition. The diatom-inferred pH profile indicates that, over the past 100 years, dramatic pH changes have occurred in Hannah Lake (Fig. 4). Prior to about 1880, the lake water pH was about 6.0 (range 5.65~.20). Since anthropogenic activity in the Sudbury area started only after 1880, these inferred pH values should be reflecting the background or "natural" pH of Hannah Lake. Soon after 1880, inferred pH started to decline from a high of ~ 6.0 to a low of ~ 4.6 in the early 1970s. The onset of the pH decline in Hannah Lake corresponds closely with the start of open pit roasting of sulfide rich ore at Copper Cliff in the 1880s (Ontario Research Foundation, 1949). The close concordance between the inferred pH decline and the increase in Cu and Ni inputs to the lake (Fig. 5) further support our hypothesis. Because the lake did not receive any mine tailings in the past, the most likely cause for the pH decline is the atmospheric deposition of strong acids. A close association between the decline in pH and the smelting activity at Copper Cliff is reasonable because of Hannah Lake's proximity (4 kin) to Copper Cliff. During the early years of ore roasting, when tall stacks were not present, the impact of atmospheric pollutants must have been restricted to the close vicinity of Copper Cliff. The pH decline was most rapid between about 1880 and 1940. Over this period, the lake's inferred pH dropped > 1 pH unit (6.0 to 4.8). Between 1940 and immediately prior to neutralization in 1975, the inferred pH declined very little

0

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12

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l 0

Composition

i 20

i

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Percent

- 1877

--

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.

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o

-

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~

4.5

; 5.5

l ~0 Diatom-inferred

t S.D

Fig. 4. The percent composition of pH-indicator diatom assemblages and inferred pH profile from Hannah Lake.

o

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2

0 -L.

o

pl-I

l 6.5

l 7.0

61 Cu and 0

1000 ~

,

~

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conc. ~ g / g d r y 2000 I

mass)

3000 ~ ~,.......~.

I

~ =": O.3

~::.: . . . . . . . . . . . .

...... ~

0.6

e- .......... ~J

0.9

/"

E

~.

/"

, ~ 1.2

1981 ~ ~ "~

•..........

.'.~ 1972 19 60 1941 o 1922

"/

f

~E 1.5

1984

"':,

'¢"~-.. " ~ " ............

...... •..........

~ tj

4000 ~

1877 ..-. . . . . . .'"

~, "'b i~ ~....'"""

1 ii t

1

2.1-i

_

/

---c~

. ....... N i

2.4

Fig. 5. Cu and Ni concentrations from Hannah Lake. (4.8 to 4.6). In 1973 and 1974, the lake's measured pH was 4.3 (Table 3) and diatoms deposited in the lake about this time indicate a lake water pH of 4.'6. The difference between measured and diatom-inferred pH is within the range of standard error of the multiple regression calibration equation ( _+0.32 pH unit), and thus we suggest that diatoms have inferred Hannah Lake's pH history with reasonable accuracy. As a result of base additions in 1975, the lake's pH increased from 4.3 to about 7.0, and since then pH has remained about 7 (Table 3). The neutralization effects are reflected in an increase in diatom-inferred pH (Fig. 4). The inferred pH increase started soon after the neutralization and the inferred pH (6.5) of the most recently deposited sediment (1984) was only 0.3 pH unit lower than the 1984 measured pH (6.8) providing additional evidence of the reliability of the lake's diatom-inferred pH history. Clearwater Lake

To infer the pH history of Clearwater Lake, the selection of a sediment core was somewhat difficult due to the difference in the distribution of surface sediment diatoms (Dixit and Evans, 1986). The predominance of planktonic diatoms in deep water surface sediments of this dimictic lake caused an overestimation of the measured pH, whereas samples from ~<8m depths provided realistic pH predictions. For these reasons, a core from 8 m depth was selected for pH reconstruction of Clearwater Lake. The core from Clearwater Lake indicated a typical log-linear relationship of

unsupported 2~°Pb activityand depth (Fig.6). Unlike Hannah Lake, there has

62 0.00

-1984

-1978 -1964

"-" - 0 . 5 0

-1949 -1930

-1910

" (J

"J

o

- 1.00

- 1890

-1.50 5 Load-210

10

Activity ( d p m / g

50 dry moss)

Fig. 6. Unsupported 2z°Pb activity in core from Clearwater Lake.

been no marked recent increase in sedimentation rate in this lake. In the past the sedimentation rate has remained ~ 13 mg cm -2 year-1, which is comparable to o t h er lakes in s out he r n Ontario (Evans and Rigler, 1983). In the core a total of 107 diatom taxa were identified (Dixit, 1986). Fourt een of these taxa were common (Fig. 7). Downcore distribution of these taxa indicates t h a t up to 1930 their populations remained unchanged, but since then major floristic changes have occurred. TabeUaria quadriseptata and Eunotia exigua were the common acidobiontic taxa. The percent abundance of these two species greatly increased in the top 2 cm of sediments. The dominance of the majority of common acidophilous taxa (all acidophilous Eunotia species, Frustulia rhomboides, Pinnularia subcapitata, P. sudetica, Surirella delicatissima, Neidiurn iridis var. amphigomphus and Navicula mediocris), also increased in post-1930 sediments. Over this period, the tax a which declined in abundance were Melosira distans, T. flocculosa, T. flocculosa IIIp, Cyclotella ocellata, and C. stelligera. Among these taxa, M. distans, T. flocculosa and T. flocculosa IIIp are acidophilous, C. ocellata is circumneutral, and C. stelligera is pH-indifferent. The five common taxa t h a t declined in abundance are all planktonic, whereas the diatoms t h a t increased in r ecent sediments are all benthic (Fig. 7). P l a n k t o n i c diatoms as an assemblage declined from a high of -,-80% to a low of ~ 15% in the surface of the core. The decline of planktonic diatoms is closely associated with the drop in lake water pH. Similarly, parallel to acidification in H a n n a h Lake, percent planktonic diatoms declined from a high of ~ 40% in about 1880 to a low of -~ 3% in 1970 (Fig. 3). This was followed by a rapid rise in planktonic diatoms after neutralization. The declines of planktonic diatoms in Clearwater and H a n n a h lakes are very similar to Big Moose Lake where planktonic diatoms declined from ~ 15 to < 1% as the lake acidified (Charles, 1984). The diatom-inferred pH profile indicates that, before 1930, the lake pH was 6.0-6.5, but since th en there was a rapid decline in pH from ~ 6.0 to ~ 4.2 (Fig.

6-

20

5

acb

5

acb

i

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i

acp

10

i

210

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{

acP

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5

i

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~1

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110

-

-

Percent

acp

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acD

10

i

__

i

Composition

Fig. 7. R e l a t i v e a b u n d a n c e of c o m m o n d i a t o m t a x a in C l e a r w a t e r L a k e .

1B"

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Composition

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4.5

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55

Diatom-inferred

6.5 pH

Fig. 8. The percent compositionof pH-indicator diatom assemblages and inferredpH profilefrom Clearwater Lake. 8). The pH decline lasted about 40 years (1930-1970). It appears that since about 1970, the lake water pH has stabilized at about pH 4.2. There was very little change in observed pH between 1973 and 1984 (Table 3), which strongly resembles the post-1970 inferred pH results. The difference between diatominferred and measured pH ( ~ 0.3 pH unit) is within the standard error of the multiple regression calibration equation. Cu 0

1

ond

Ni c o n c .

(.ug/g

800 t

t

dry

,~

mass)

1000 t

t .* 2'

o5

,

f/~ ....., .............. ./.:'.....

~.o

~E o ]£

1.5

1500 1984 1978

,9,4

1949 .~ ~93o

IE!¢" .~.:

1910

~

1890

20 ~

---cu

......... Ni 2.5

Fig. 9. Cu and Ni concentrations from Clearwater Lake.

65 In Clearwater Lake the concentrations of Cu and Ni, prior to anthropogenic activity (Fig. 9), are comparable to background concentrations in lakes from several regions of North America. However, at the beginning of this century C u and Ni inputs increased causing the observed changes in sediment concentrations. By the end of the 1920s the metal inputs had increased, but the largest increase occurred after 1920 when the installation of a tall stack at Copper Cliff and the Falconbridge Smelter were completed. In Clearwater Lake, acidification started about 50 years later than in Hannah Lake. Although the difference in natural buffering ability of lakes and their watersheds may be responsible for this delay, the distance of lakes from Copper Cliff and the installation of tall stacks in the 1920s are likely to have played an important role. Clearwater Lake is located 12km downwind from Copper Cliff, whereas Hannah Lake is only 4 km away. In the later part of the last century, when open pit roasting started at Copper Cliff, the influence of atmospheric gases must have been restricted to nearby lakes (e.g. Hannah). The open pit roasting was discontinued around 1923 when the installation of a 155 m stack and sulfur recovery plant was completed at Copper Cliff (Nriagu et al., 1982). The operation of the stack may have lowered the influence of atmospheric pollutants in the immediate vicinity, but increased the effect on lakes located further from the source. After the installation of the superstack in 1972, SO2 emissions have significantly declined in the Sudbury area (Dillon et al., 1986). The reduction in SO2 emission appears to be the most likely cause for the cessation of any further pH decline in Clearwater Lake over the last decade. CONCLUSIONS Stratigraphic analysis of diatoms from Hannah and Clearwater lakes provides useful information of the past lake water pH changes in these two lakes. A close correspondence between the diatom-inferred and observed pH measurements made since 1969 suggests that the pH changes inferred by this study are real and accurate. As a result of increased SO2 emissions from the mining and smelting activities in Sudbury, gradual acidification has occurred in lakes around Sudbury. In lakes located closest to Copper Cliff, the pH decline commenced soon after the roasting of ore started, whereas in lakes located away from the source, the acidification seems to have started after the installation of tall stacks in the 1920s. Diatom-inferred and measured pH results of Clearwater Lake suggest that since the early 1970s any further pH decline has halted. This is consistent with the reduction in SO2 emissions after the installation of the 381 m superstack in 1972. ACKNOWLEDGEMENTS This paper is a part of the Ph.D. thesis of S.S. Dixit submitted to Queen's University as part of the requirements of the Trent-Queen's Graduate Program. Financial support was made available in the form of a Queen's Graduate

66

Fellowship to S.S.D. and a research grant from the Natural Sciences and Engineering Research Council of Canada to R.D.E. Thanks are due to Kevin Kinney for doing the 21°pb and metal analyses and to the Ontario Ministry of the Environment for providing the unpublished lake pH data. REFERENCES Anderson, D.S., R.B. Davis and F. Berge, 1986. Relationships between diatom assemblages in lake surface-sediments and limnological characteristics in southern Norway. In: J.P. Stool, R.W. Battarbee, R.B. Davis and J. Meril~iinen (Eds), Diatoms and Lake Acidity. Junk, Dordrecht, The Netherlands, pp. 97-104. Appleby, P.G. and F. Oldfield, 1978. The calculation of 21°pb dates assuming a constant rate of supply of unsupported 21°pb to the sediment. Catena, 5: 1~. Battarbee, R.W., 1973. A new method for the estimation of absolute microfossil numbers, with special reference to diatoms. Limnol. Oceanogr., 18: 647~53. Bronson, R.R., 1975. Evaluation of a rapid procedure for heavy metal analysis in sediment and soil. Work Report, University of Waterloo, Waterloo, Ontario. Chan, W.H., J.S. Tang and M.A. Lusis, 1982. An analysis of the impact of smelter emissions on precipitation quality and wet deposition in the Sudbury area: Sudbury environmental study event precipitation network results. Ont. Min. Environ. Rep. SES 006]82, 72 pp. Charles, D.F., 1984. Recent pH history of Big Moose Lake (Adirondack Mountains, New York, USA) inferred from sediment diatom assemblages. Verh. Int. Ver. Limnol., 22: 559-566. Charles, D.F., 1985. Relationships between surface sediment diatom assemblages and lake water characteristics in Aridondack lakes. Ecology, 66: 994-1011. Cleve-Euler, A., 1951-55. Die Diatomeen von Schweden und Finland. K. sv. Vet. Akad. Ser. IV: 961 pp. Davis, R.B., S.A. Norton, C.T. Hess and D.F. Brakke, 1983. Paleolimnological reconstruction of the effects of atmospheric deposition of acids and heavy metals on the chemistry and biology of lakes in New England and Norway. Hydrobiologia, 103: 113-123. Dillon, P.J., R.A. Reid and R. Girard, 1986. Changes in the chemistry of lakes near Sudbury, Ontario following reductions of SO2 emissions. Water Air Soil Pollut., 31: 59-66. Dixit, S.S., 1986. Algal microfossils and geochemical reconstructions of Sudbury lakes: a test of the paleoindicator potential of diatoms and chrysophytes. Ph.D. Thesis, Queen's University, Kingston, Ontario, 190 pp. Dixit, S.S. and M.D. Dickman, 1986. Correlation of surface sediment diatoms with the present lake water pH in 28 Algoma lakes, Ontario, Canada. Hydrobiologia, 131: 133-143. Dixit, S.S. and R.D. Evans, 1986. Spatial variability in sedimentary algal microfossils and its bearing on diatom-inferred pH. Can. J. Fish Aquat. Sci., 43: 1836-1845. Eakins, J.D. and R.I. Morrison, 1978. A new procedure for the determination of lead-210 in lake and marine sediments. Int. J. Appl. Radiat. Isot., 29: 531-536. Evans, H.E. and D.C. Lasenby, 1984. A comparison of lead and zinc sediment profiles taken by diver and a gravity corer. Hydrobiologia, 108: 165-169. Evans, R.D., 1980. Measurement of sediment accumulation and phosphorus retention using lead210 dating. Ph.D. Thesis, McGill University, Montreal, Quebec, 169 pp. Evans, R.D. and F.H. Rigler, 1983. A test of lead-210 dating for the measurement of whole-lake soft sediment accumulation. Can. J. Fish. Aquat. Sci., 40: 7~81. Flower, R.J. and R.W. Battarbee, 1983. Diatom evidence of recent acidification of two Scottish lakes. Nature, 305: 130-133. Foged, N., 1981. Diatoms in Alaska. J. Cramer, Vaduz, Germany, 317 pp. Germain, H., 1981. Flore des Diatom~es. N. Boub~e, Paris, 444 pp. Gorham, E. and A.G. Gordon, 1960. The influence of smelter fumes upon the chemical composition of lake waters near Sudbury, Ontario, and upon the surrounding vegetation. Can. J. Bot., 38: 477-487.

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