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Richness of aquatic macrophyte floras of soft water lakes of differing pH and trace metal content in Ontario, Canada
Aquatic Botany, 23 ( 1 9 8 5 ) 27--40
27
Elsevier Science Publishers B.V., A m s t e r d a m -- Printed in The N e t h e r l a n d s
RICHNESS OF AQUATIC MACROPHYTE FLORAS OF SOFT WATER LAKES OF DIFFERING pH AND TRACE METAL CONTENT IN ONTARIO, CANADA
N.D. Y A N , G.E. M I L L E R I, I. WILE s and G.G. H I T C H I N g
Ontario Ministry o f the Environment, Dorset Research Centre, P.O. Box 39, Dorset, Ont. POA lEO (Canada) ( A c c e p t e d for p u b l i c a t i o n 16 May 1985)
ABSTRACT Yan, N.D., Miller, G.E., Wile, I. and Hitchin, G.G., 1985. Richness o f aquatic m a c r o p h y t e floras o f soft w a t e r lakes of differing pH and trace m e t a l c o n t e n t in Ontario, Canada. Aquat. Bot., 23: 27--40. The richness of t h e aquatic m a c r o p h y t e floras, i.e., the total n u m b e r o f species, was assessed in 39 soft w a t e r lakes in central Ontario, Canada. T h e Cu and Ni c o n c e n t r a t i o n s and pH o f the lakes ranged f r o m 1 to 360 mg m -~, 2 to 3 7 0 0 mg m -3 and 3.9 to 7.0, respectively. T w o non-exclusive subsets of the data were e x a m i n e d to d e t e r m i n e firstly, if floral richness was related to lake pH in lakes with low Cu and Ni levels (Data Set I) and secondly, if floral richness in acidic (pH < 5.3) lakes was related to levels o f various trace metals (Data Set II). C h a r o p h y t e s were not f o u n d in lakes with pH < 5.2. In Data Set I, there was no relationship b e t w e e n the richness of t r a c h e o p h y t e s and pH, and there was a negative relationship b e t w e e n pH and b r y o p h y t e richness. Unlike p h y t o p l a n k t o n , z o o p l a n k t o n , b e n t h i c m a c r o i n v e r t e b r a t e s and fish, there was no decrease in t o t a l species richness in lakes o f pH < 5.5, as long as trace m e t a l levels were low. E x a m i n a t i o n o f Data Set II indicated t r a c h e o p h y t e richness of acidic lakes was negatively correlated w i t h Cu and Ni levels. Biological surveys of m e t a l - c o n t a m i n a t e d acidic lakes are, t h e r e f o r e , n o t of use for predicting the effects of acid deposition alone on aquatic m a c r o p h y t e s .
INTRODUCTION
Much of our information concerning the effects of acid deposition on aquatic biota has been derived from surveys of large numbers of lakes distributed along a pH gradient. Such surveys have demonstrated that the number of species of p h y t o p l a n k t o n (Almer et al., 1974; Leivestad et al., 1976), zooplankton (Sprules, 1975; Almer et al., 1978), benthic macroinvertebrates (Hendrey and Wright, 1976; Okland and q)kland, 1980) and Ontario Ministry o f the E n v i r o n m e n t , 83 A l g o n q u i n Blvd. W., T i m m i n s , Ont. P4N 2R4 Canada. Ontario Ministry of the E n v i r o n m e n t , 40 St. Clair Ave. W., T o r o n t o , Ont. M 4 V 1M2 Canada. 3 Ontario Ministry o f the E n v i r o n m e n t , P.O. Box 213, Rexdale, Ont. M9W 5L1, Canada.
0304-3770/8.5/$03.30
© 1985 Elsevier Science Publishers B.V.
28
fish (Harvey, 1975; Almer et al., 1978) is reduced when pH levels o f lake waters fall much below 6.0. While intensive surveys of small numbers o f lakes in Scandinavia have demonstrated that acidification may alter macrop h y t e assemblages (Grahn, 1977; Nilssen, 1980), the richness of the floras has rarely been assessed for a large number o f lakes. Hence, it is not yet k n o w n if the richness of macrophyte floras is positively correlated with lake pH, as is the richness of all other groups of aquatic biota. Roelofs (1983, and personal communication, 1984) indicated the richness of aquatic m a c r o p h y t e floras of lakes in The Netherlands has been reduced by acidification. However, the average pH o f these lakes is 3.9 (Roelofs et al., 1984). Lakes with pH < 4.5 are very rare in Precambrian Shield areas o f Europe and North America, where acid deposition rates are elevated (Wright et al., 1980). There has apparently been only one survey reporting the richness of macrophyte floras of r e m o t e recently-acidified lakes with a more representative range in pH. Roberts et al. (1985) surveyed 9 clear water lakes in the Adirondack Mountains of New York which ranged in pH from 4.4 to 6.9. Richness was lower in the acidified lakes, although Roberts and colleagues felt their data set was t o o small to ascribe this trend to acidity. The first purpose of this report is to determine whether macrop h y t e richness (the number of distinguishable taxa per lake) is also lower in acidified lakes in Ontario, Canada. The interpretation of results o f biological surveys of acidified lakes is frequently complicated b y the additional contamination of the acidic lakes with potentially toxic trace metals discharged from local point sources (e.g., Gorham and Gordon, 1963; Yan and Strus, 1980; Havas and Hutchinson, 1983). The second purpose of this report is to investigate whether the richness o f the m a c r o p h y t e floras of acidified lakes might be reduced b y their contamination with potentially toxic trace metals, Cu and Ni in particular, metals which unlike A1 and Mn (Wright et al., 1980) are not normally found at elevated levels in acidified lakes. METHODS
Miller et al. (1983) measured trace metal levels in tissues of aquatic macrophytes collected from 46 lakes in Ontario, ranging in pH from 3.9 to 7.0. We assembled m a c r o p h y t e richness data from 39 o f these lakes. Six of the lakes were excluded because their chemistry had been manipulated for experimental or other reasons (Hitchin et al., 1984); the seventh was excluded from our data sets (to be discussed) b y virtue o f its Ni level (> 10 mg m -3) and pH value (> 5.3). Sixteen o f the 39 lakes were located in the Sudbury area, the remaining 24, hereafter called the " D o r s e t " lakes, in south-central Ontario (Fig. 1). Each lake was visited once between May and September (usually July or August) of 1977, 1978 or 1979. Divers using snorkeling or SCUBA gear first assessed the heterogeneity of plant distribution in each lake. Consider-
29
• ,
ONTARIO
I I
,
QUEBEC
i
l
÷ O SUDBUR
°~°~< 9
oO
°
+
.:
•
0
I00 krn
._Z Fig. 1. L o c a t i o n o f s t u d y lakes. Also indicated is the z o n e within which lake p H is generally < 5.5 (cross-hatched), and d o t t e d isoline within which surface w a t e r Ni levels e x c e e d 100 mg m -3 ( f r o m Anon., 1978).
ing lake size and the results of the preliminary survey, a variable n u m b e r of sites (usually 5--15) was selected for detailed inspection. At each site, divers identified all species of submerged and floating-leaved aquatic tracheophytes, b r y o p h y t e s and charophytes found on a 10 m-wide transect extending from the shore to the maximum depth o f plant occurrence. B r y o p h y t e identifications were confirmed b y C. Manville (Royal Ontario Museum, Toronto). Macrophyte richness was not correlated with numbers of transects examined (Hitchin et al., 1984). Surface lakewater pH was measured in the field with a Radiometer Model 29 pH meter in 1--3 littoral areas, depending on lake size. The pH was measured in situ after appropriate buffer calibrations and after the readings had stabilized (< 5 min in all cases). Trace metal concentrations of surficial littoral sediments and epilimnetic or whole lake composite samples were determined as described b y Miller et al. (1983) and Hitchin et al. (1984), respectively. Divers classified each lake as highly, or n o t highly coloured (see Fig. 2).
30 TABLE
1
Data set assignment, Ni levels, pH and aquatic macrophytes
of the study lakes
x
o
~~ Data set
Ni (mg m
pH
Lake
11
3700
4.7
54*
taxa
~)
3.9
14'
300
3.9
Swal~*
*
4
4 x
II II
260
4.4
Clearwater
8
x
120
5.1
4*
7
x
I1
85
5.2
75*
9
x
l[
22
4.5
Terry
17
nd
x
x
x
Freela
17
x
x
Bell
32
x
x
x
x x
x x x
x
x
x
x x x
x
4.8
George
23
x
x
x
II
6
4.6
Carlyle
30
x
x
x
II II II II
4 ~2 -3 -:2
4.0 4.2 5.2 5.3
Krarner * * Horn" plastic Cinder
5 17 17 15
x
11
,:2
5.3
Crosson
19
x
x
II
<2
4.4
Axe
24
x
x
x
11
-2
5.3
Fawn
24
x
x
x
x x x
5.5
Leonard
16
x
x
x
,4
5.5
Heney
28
x
x
x
x
<~3
5.6
MeKay
27
x
x
x
42
5.6
Healey*
22
x
<3 <2 <~2 2
5.7 5.7 5.8 5.9
Leech 94* Gud feather ÷ Moot*
17 22 19 15
2
5.0
x x
x x
x x
x
x x
x x
x x x
x
x x
x x
x x x
x
x x
x x x
x
x x x
x x
x
x
x x x
x x
x
x
x
x x x
x
x
x x x x
x x x x
x
x
x
x x x
x
x x x
x
x
x
x x
Hillman +
19
x
x
6.4
Clear
l 5
x
x
x
6.4 6.4 6.5
Brandy Otter Little Otter R e d Chalk Frond Harp LittLe C l e a r Solitaire
18 27 25 12
x
x x x
x x x
44
x
x
x x
codes
follow
**Local name only. ***Ni ieveB una~ilable eoloured
Gorham but
lakes.
assumed
and
Gordon tt) b~
x
x
x
x
x
x
x
x x x
x
x
x x
x
x
x
x
x
x
x
x
x
x x
x
<']0
x
(I 9 6 3 } , mR
m
' based
on
hike
location.
x
x x x
x
x
x
x
x x
x
x x
x x x x x x
x x
x
x
x
x
x
x x
x
x
x x x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x x x
x
x
x
x
x
x
x
x x
x
x
x
x
x
x x
x
x x
x
x x
x
x
x
x x
x
x x x
x
x x x
x
x
x x x
x
x X
x
x x x
x
x
x
x x
x x
x
x x
x x
x
x x
x
x x
x
x
x
x
x x
x
x x x
x
x
x x
x x
x
x
x x
x x
x
x
x x
x
x x
x x x
x x x x x
x
6.3
x
x x
x
x x
x
x x
x x
x x
x
x x
X X
x x
x
I1 16 6
x
x
18
6.8 6.9 7.0
x x
19
<2 "~2 < 2
x
x x x x
2]
x
x x
103"
x
x x
x x
Dickie
**
x
x x
Chub
6.6 6.6
x
x
x x
x x x
x x
6.0
--*
* Highly
x
x
x x x
x x x
x
x
x
x
x x
x x
x x
x
x x x
6.2
2 --** * --*** 2
o
x
x x x
x x
3
< 2
x x
x
x x
• 2
--***
*Numeric
x
x ×
x x
fi
<
o
x x
II
3
o
x
x
x
4.8
x
x x
x
4.3
+
o
x
660
1l
0
x
x
[[ II
• :20
0
~
3
÷
o
x
x
x x x
x
x
~ x x
x
~x x
x
x
x
x
~t
xx~
Sull.
P. B e a u v . a m b l y p h y l l u r a
Spha~um
tere~ ( S c h i r a p . ) ~ t n g s t r .
Sphagnum sub~ecundum N c e s & H o r n s c h .
Sphagnum r e c u r u u m
Sphagnum pylae|ii Brld.
Sphagnum palu#tre L.
Sphagnum mqJua ( R u s s . ) C, Jcns.
Sphagnum ftmbr~atum W i l s .
Sphagnum cuJpidatum E h r h .
(Row.)Warnst.
Pohlia nu tanm ( H e d w .) L i n d b . s c h i m p e r i i ( H e d w .) L i n d b .
Hygroomblymtegium tenox ( H e d w . ) Jenn.
Gymnocolea inflata ( H u d s . ) D u m .
Fontinal~ nova-angliae
Fontinalis durlaei S c h i m p .
Fonttnalil hypnotde# R . H a r t m a n n
FontinoJim anttpyrettca H e d w .
Fis#iden# fort tanuJ ( B . P y l . ) S t e u d .
Drepanocladus sp.
CIadopodiella nultanw (Nees) Buch.
Nitella sp,
d u r i a e i F.
Bruz.) Ag.
Nite~ia tenuluima ( D e s v . ) K / / t ; ' .
Nitella gracili~ (Sin,) A g ,
Nitella furcata ( R o x h . e x
Nitella flexilis ( L . ) A g .
Chara sp.
ValllzneMa ame~cana M i c h x
UO'icular4a uutgarim L.
Utr~cularia re#uplnata B . D . G r e e n
Utricularta purpurea W a l t ,
Utricu~aria m i n o r L.
Utrtcular~a intermedi= H a y n e
Utr~culari* gtbba L.
Utrtcularla c o ~ u t a M i c h x
Sparganium sp.
Sagittar~a sp.
Ranuneulu~ ~ep Fans L.
Potamogeton uo,w#eyl R o b b i n s
Potamogeton #pirillbl T u c k e r m .
Potamogeton robbln#ff O a k e s
Potomageton Hchard#onil ( B e r n . ) Rydb.
Potaenogeton obtu,ifoltus M e r t . & K o c h . PotamoEeton puMIlus L.
O0 b.a
32
Description o f the study lakes The pH and Ni levels o f the study lakes are listed in Table I. Lake pH values ranged from 3.9 to 6.6 among the Sudbury lakes and from 4.2 to 7.0 among the Dorset lakes. Several o f the acidic Dorset lakes had coloured waters, indicating their acidity was partially attributable to organic acids. Levels of Ca ranged from 2 to 5 mg 1-1 among the Dorset lakes and from 2 to 15 mg 1-1 among the Sudbury lakes. The elevated levels were f o u n d in the acidic lakes near Sudbury and were attributable to elevated rates o f weathering of watershed materials produced by input of acids (Dillon, 1983). Levels of Cu and Ni were greatly elevated in lakes within 25 km o f the Sudbury smelting complex (Fig. 1 and Conroy et al., 1975). Levels were much lower in the acidic lakes located further from the city and in all lakes with pH values > 5.5. Eriocaulon septangulare With. and Eleocharis acicularis (L.) R. and S. were the most frequently encountered species, occurring in over 75% of the lakes. An additional 14 species, occurred in half or more t h a n half o f the lakes (Table I). Roberts et al. (1985) identified Potamogeton confervoides Reichenb. and Sphagnum spp. as characteristic o f acidic lakes in the Adirondacks. This agrees with their occurrence in the Ontario study lakes. P. confervoides was f o u n d in 9 lakes with a median pH of 4.8 (range 4.3--5.6). Sphagnum spp. were f o u n d in 15 lakes ranging in pH from 4.2 to 6.4 with a median pH of 5.3. While various Utricularia species were observed in the study lakes, Utricularia geminiscapa Benj. was not observed. This species is characteristic o f acidic lakes in the Adirondacks (Roberts et al., 1985). The plant assemblages of the study lakes are described in detail elsewhere (Wile and Miller, 1983). RESULTS
The survey data were organized into 2 subsets to address the 2 stated purposes. To examine the influence o f depressed pH itself on m a c r o p h y t e richness, data from those 30 lakes with lake water Ni levels < 10 mg m -3 were examined (Data Set 1). As Ni levels in lake water were positively correlated (P < 0.01) with levels of Ni (r = 0.95) and Cu (r = 0.91) in littoral sediments and with Cu levels in lake water (r -- 0.98, n = 40), the Ni levels in lake water were assumed to provide a satisfactory index o f Ni and Cu contamination of waters b o t h above and below the sediment surface. Copper can be extracted both from the water and the sediments by macrophytes (Peter et al., 1979). The 18 lakes with pH ~< 5.3 were operationally termed " a c i d i c " and formed Data Set 2. This set o f data was assembled to examine the relationship between m a c r o p h y t e richness in acidic lakes and habitat metal contamination. A separatory pH of 5.3 was chosen for 2 reasons. It approxi-
33
mates the equivalence point o f alkalinity titrations of soft waters; hence, soft water lakes with pH ~< 5.3 should have measurable free acidity ( S t u m m and Morgan, 1981). Secondly, changes in macrophyte c o m m u n i t y structure have been observed at approximately this pH (Roberts et al., 1985). By design, ranges o f Cu and Ni levels in lake water were low in Data Set 1 (1--3 mg m -3 and 2--6 mg m -3, respectively), b u t ranges of levels o f o t h e r trace metals were quite large (3--210 mg Zn m -3, 18--135 mg Mn m -3, 30--1150 mg Fe m -3 and 10--360 mg A1 m-3). The lakes in Data Set 1 ranged in pH from 4.0 to 7.0, and supported an average of 20.1 (range 5--44) species of aquatic macrophytes, o f which 16.9 (4--39) were tracheophytes, 2.4 (0--6) were b r y o p h y t e s and 0.8 (0--3) were charophytes. Significantly more species of aquatic b r y o p h y t e s (Musci) were found in the acidic lakes (Table II, Fig. 2). A positive correlation between b r y o p h y t e richness and lake area was also observed. This was n o t an indirect consequence o f a lake area--lake pH relationship (r = 0.26, P 0.10). It m a y simply reflect increased habitat area in larger lakes. Nevertheless, a greater relative and absolute richness of species o f the k n o w n acidophilous taxon, Sphagnum, in lakes of pH < 5.0 (Table I and Wile and Miller, 1983) suggests that it is probably lake acidity that directly or indirectly controls b r y o p h y t e richness. TABLE II Correlation coefficients between aquatic plant richness and selected limnological variables for lakes in Data Set 1. Significance at 95 and 99% indicated by single and double asterisks, respectively Parameter
Bryophyte
pH --0.47** A1 --0.05 Ca --0.21 Conductivity --0.11 Lake area 0.38*
Charophyte
Tracheophyte
Macrophyte
0.38* --0.11 0.50** 0.63" * 0.03
0.01 --0.28 0.60** 0.31 0.21
--0.04 --0.24 0.55** 0.32 0.25
Charophytes were observed in 15 lakes, all with pH > 5.2 (Fig. 2). In his review, Hutchinson (1975) noted that the majority of charophyte species are found in alkaline waters. The positive correlations between c h a r o p h y t e richness and pH, Ca and conductivity (Table II) are consistent with this observation. The richness of aquatic tracheophytes was not significantly correlated with pH, conductivity or lake area in Data Set 1 (Table II). Additionally there was not a significant correlation with A1, the toxic metal p r o b a b l y directly responsible for lowering the richness of fish communities in acid lakes (Baker and Schofield, 1982). Richness was correlated with Ca levels. This was in part attributable to the unusually high richness o f F r o o d Lake
34 L~ < w I DO n~ •
I C)
•
•
•
•
~
•
o
•
•
8 o• co w Z I
4 •
2
o0
2
•
e g o •
:
• g o
• •
0
EE
v
~
i
1
T
•
~
o
oo
i
T
w >I o_ 0 EK
40
<
~:
3o
o I u
<
o
20
~ •
OE I--
•
0 00
•
•
0
•
•
0
O
•
•
IO-
0
i
4.0
i
4.5
i
5.0
i
5.5
i
i
i
i
6.0
6.5
7.0
7.5
pH Fig. 2. Scattergrarns of pH vs. numbers o f species o f aquatic Tracheophyta, Musci and Charophyceae in each lake in Data Set 1. Highly coloured lakes are indicated by open circles for the Tracheophyta. Numbers indicate the number o f coincident points.
(39 taxa), the lake with the highest Ca levels. However, elimination of Frood Lake from the data matrix did not eliminate the correlation o f tracheophyte richness with Ca (r = 0.40, P < 0.05). There were 9 coloured lakes in Data Set 1. The pH o f these lakes ranged from 4.2 to 6.3. The tracheophyte richness o f the coloured lakes was similar to that of the lakes with greater transparency (Fig. 2), indicating that tracheophyte richness in recently acidified, clear water lakes is apparently not different from acidic lakes with coloured waters. Such lakes presumably have been acidic for longer periods o f time. There may, o f course, be differences in species assemblages between the coloured and clear water acidic
35 lakes; however, as such differences were not detected in the ordination of species presence--absence data {Wile and Miller, 1983), any differences must be small. Because of the opposing patterns of bryophyte and charophyte richness and because of the poor correlation between tracheophyte richness and 30-
30-
20
•
20-
I0
~0
0o
. . . . . . . . . . .~. . . . . .
S o
--
[Fe] w Z T
F-)T 13O W I (J
........
O
........
o
b
-
O
-
.......
;
........
£
b
g
9
~n]
( m g m -3 )
30
30-
20
20
10
I0
0 o
v
0
(mg m -3)
. . . . . . . .
0 o
Or"
~D 8
(mg m "3) 30-
30
20
2
20
:
•
10
I0
•
0
. . . . . . .
o
-
. . . . . . .
o
o
B . . . . . . .
• ~
o
o
•
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
-
o
o
o°
o°
Fig. 3. Scattergrams of tracheophyte richness vs. water column levels of Fe, Zn, AI, Mn, C u a n d Ni for lakes in Data Set 2. Three values were unavailable for the Al plot.
36 pH, there was no relationship between total m a c r o p h y t e richness and pH (r = --0.04, P > 0.10). The analysis of Data Set 1 produced significant correlations b e t w e e n pH and the richness of b r y o p h y t e and charophyte, but n o t tracheophyte floras. Therefore, we chose to examine only t r a c h e o p h y t e data in Data Set 2, the set o f all lakes with pH ~< 5.3. Lake pH varied from 4.0 to 5.3 for the 18 lakes in Data Set 2 (Table I) and was n o t significantly correlated (r = 0.35, P > 0.10) with tracheophyte richness. Ranges o f Zn, Mn, Fe and A1 were similar in this data set to those in Data Set 1 (3--210, 43--330, 30--1150 and 27--360 mg m -3, respectively). Ranges of Cu (1--360 mg m -3) and Ni (2--3700 mg m-3), were, of course, much wider. Scattergrams o f tracheophyte richness and lakewater levels o f Fe, Zn, Mn, A1, Cu and Ni are presented in Fig. 3. No pattern is evident for Fe and Zn. For A1 and Mn, there is some indication o f reduced richness at the highest metal levels. There is a clear pattern o f decreasing richness with increasing levels of Cu and Ni. DISCUSSION
Acidification and macrophyte richness Summarizing the results of various surveys o f naturally acidic lakes, Hutchinson (1975) indicated that charophytes are generally not observed in lakes of pH < 5.5. Therefore, the absence of charophytes from lakes of pH < 5.2 in this survey (Fig. 2) is consistent with k n o w n patterns of occurrence of the group. As charophytes similarly have n o t appeared on any m a c r o p h y t e species lists from aquatic plant surveys of recently acidified lakes (Hendry and Vertucci, 1980; Eriksson et al., 1983; Roelofs, 1983; R o b e r t s et al., 1985), it may be postulated that the charophycean flora of soft water lakes will probably be eliminated should lake pH fall to 5.0 or below. With the exception of the charophytes, which made only a small contribution to total macrophyte richness, there was no evidence of reduced m a c r o p h y t e richness in lakes of low pH (Fig. 2). Therefore, in terms o f richness, m a c r o p h y t e s differ from plankton, benthic macroinvertebrates and fish, all o f which have reduced richness in acidic lakes (Dillon et al., 1984). Most of the m a c r o p h y t e species found in these s t u d y lakes actually occurred over the entire pH range of 4.0--7.0 (Table I). While this seems to indicate that most tracheophyte species observed in soft water Shield lakes are tolerant of substantial pH depressions, their tolerance may be more apparent than actual. The roots o f aquatic m a c r o p h y t e s of soft water lakes are the primary site of uptake and exchange of nutrients and major ions (Wium-Andersen, 1971; Carignan and Kalff, 1980). Below the t o p few centimetres, the pH o f the porewater of littoral zone sediments in acidic lakes is generally much higher than that o f the overlying water (Hultberg
37
and Grahn, 1975; Roberts et al., 1985). Hence, most exchange o f materials in these m a c r o p h y t e s probably takes place across membranes which are n o t exposed to strongly acidic water. Collins et al. (1981) similarly indicated that the isolation of benthic invertebrates found beneath the surface o f the sediments o f lakes (infauna) from water of lowest pH might explain w h y these organisms are n o t as severely affected b y the acidification o f lakes as epifaunal invertebrates. It is n o t surprising that acidification has reduced the richness o f lakes in The Netherlands, b u t not in Ontario. With an average pH of 3.9, those lakes studied b y Roelofs et al. {1984) are more acidic than the lakes studied herein. Further, the Dutch lakes have p r o b a b l y been acidic for longer periods of time (Beamish and Harvey, 1972; Roelofs, 1983). As acidification of sediment pore water lags behind that of lake waters (Oliver and Kelso, 1983) and as acidification is probably a more recent p h e n o m e n o n in Ontario, it may be that detrimental impacts on the m a c r o p h y t e floras of Ontario lakes are yet to be observed. Roberts et al. (1985) surveyed 9 lakes in the Adirondack Mountains of New York which ranged in pH from 4.4 to 6.9 and noted that total macrophyte richness was lower in the acidic than in the non-acidic lakes. In contrast, results of our survey indicate that the depressions in pH of the magnitude o f that usually reported in Canadian Shield Lakes (pH > 4.5, Wright et al., 1980) probably will n o t result in reductions in aquatic b r y o p h y t e or tracheophyte richness. Intensive m a c r o p h y t e surveys have indicated that acidification m a y alter the biomass (Wile et al., 1985) and composition of m a c r o p h y t e communities (Hultberg and Grahn, 1975; Roelofs, 1983) and the distribution of macrop h y t e species with depth (Singer et al., 1983). However, for lakes in Ontario, it does n o t appear that acidification of lakes has reduced macrophyte richness.
Trace metal contamination o f acidic lakes and macrophyte richness There was no relationship between Fe and Zn levels and tracheophyte richness among the acidic lakes (Fig. 3). Relationships were not anticipated with Fe as it is not a particularly p h y t o t o x i c element (Stokes, 1983) and as the observed Fe concentrations were n o t unusually high for Shield lakes. Concentrations of Zn were much higher in some of the acidic lakes in Data Set 2 than typically observed in either acidic or non-acidic Shield lakes (Borg, 1983). Nevertheless, toxic effects of Zn on macrophytes are not evident until levels are orders of magnitude higher than Zn levels recorded in this study (Stanley, 1974). Hence, a relationship between Zn levels and t r a c h e o p h y t e richness was not anticipated. Tracheophyte richness was low in lakes with high levels of Cu and Ni, supporting Gotham and Gordon's (1963) hypothesis that their observed positive relationship between m a c r o p h y t e richness and distance of lakes
38
from Sudbury might be related to levels of Cu and Ni. Tracheophyte richness was also reduced in a few lakes with approximately 300 mg m -3 or more o f A1 and 150 mg m -3 or more o f Mn (Fig. 3). It is, o f course, not possible to use these survey data to identify the cause of the unusually low tracheop h y t e richness in the metal-contaminated acidic lakes. However, if metal contamination is the cause, there are three reasons w h y Cu and Ni are more likely causative agents than A1 or Mn. Firstly, Scandinavian acidic lakes o f t e n have levels of A1 and Mn similar to those observed in the acidic lakes in Data Set 2 (Almer et al., 1978; Wright et al., 1980; Borg, 1983), b u t t h e y do not have elevated levels o f Cu and Ni, nor do they have unusually low m a c r o p h y t e richness (Eriksson et al., 1983). Secondly, Cu is much more p h y t o t o x i c than A1 and Mn (Stokes, 1983; Stanley, 1974) and Ni can seriously damage aquatic macrophytes at levels observed in Data Set 2, a b o u t 300 mg m -3 (Hutchinson and Czyrska, 1975; Brown and Rattigan, 1979). Finally, as Hutchinson and Czyrska (1975) reported synergism in the toxicity of Cu and Ni to aquatic plants, the presence o f high concentrations o f b o t h metals in several lakes in Data Set 2 probably exacerbates their toxic effects on macrophytes. Biological surveys o f acidic freshwaters have o f t e n included lakes or streams which are both acidic and contaminated with trace metals attributable to local industrial operations (Yan and Strus, 1980) to natural fumigations with metal-laden particulates (Havas and Hutchinson, 1983) or to drainage from acidic mine tailings (Hargreaves et al., 1975). This study indicates that the results of such surveys m a y n o t be o f use in predicting the consequences for aquatic biota of excessive inputs of strong acids alone. ACKNOWLEDGEMENTS
Appreciation is extended to W. Kerr, J. Scholer, S. Painter and D. Snell for assistance in the field and laboratory, to J. Roelofs for advice on use of his data and to B. Barley, R. Hall, D. R o b e r t s and P. Stokes for fruitful discussions and reviews of early drafts o f the manuscript.
REFERENCES Almer, B., Dickson, W., Ekstr6m, C., H6rnstr6m, E. and Miller, U., 1974. Effects of acidification on Swedish lakes. Ambio, 3: 30--36. Almer, B., Dickson, W., Ekstr6m, C. and H6rnstr6m, E., 1978. Sulfur pollution and the aquatic ecosystem. In: J.O. Nriagu (Editor), Sulfur in the Environment: Part II Ecological Impacts. Wiley, New York, pp. 271--311. Anonymous, 1978. Extensive monitoring of lakes in the greater Sudbury area. Ontario Min. Environ. Rep., 40 pp. Baker, J.P. and Schofield, C.L., 1982. Aluminum toxicity to fish in acidic waters. Water, Air, Soil Pollut., 6: 501--514. Beamish, R.J. and Harvey, H.H., 1972. Acidification of the La Cloche Mountain Lakes, Ontario, and resulting fish mortalities. J. Fish. Res. Bd. Canada, 29: 1131--1143.
39 Borg, H., 1983. Trace metals in Swedish natural freshwaters. Hydrobiologia, 101: 27--34. Brown, B.T. and Rattigan, B.M., 1979. Toxicity of soluble copper and other metal ions to Elodea canadensis. Environ. Pollut., 20: 303--314. Carignan, R. and Kalff, J., 1980. Phosphorus sources for aquatic weeds: water or sediments? Science, 207: 987--989. Collins, N.C., Zimmerman, A.P. and Knoechel, R., 1981. Comparisons of benthic infauna and epifauna biomasses in acidified and nonacidified Ontario lakes. In: R. Singer (Editor), Effects of Acid Precipitation on Benthos. North Am. Benthol. Soc., Hamilton, New York, pp. 35--48. Conroy, N., Hawley, K., Keller, W. and Lafrance, C., 1975. The influence of the atmosphere on lakes in the Sudbury area. Proc. First Spec. Symp. Atmospheric Contribution to the Chemistry of Lake Waters, Int. Assoc. Great Lakes Res., Sep. 28--Oct. 1, 1975, pp. 146--165. Dillon, P.J., 1983. Chemical alterations of surface waters by acidic deposition in Canada. Water Qual. Bull., 8: 127--132. Dillon, P.J., Yan, N.D. and Harvey, H.H., 1984. Acidic deposition: effects on aquatic ecosystems. CRC Crit. Rev. Environ. Control, 13: 167--194. Eriksson, F., H6mstrSm, E., Mossberg, P. and Nyberg, P., 1983. Ecological effects of lime treatment of acidified lakes and rivers in Sweden. Hydrobiologia, 101: 145--164. Gorham, E. and Gordon, A.G., 1963. Some effects of smelter pollution u p o n aquatic vegetation near Sudbury, Ontario. Can. J. Bot., 41: 371--378. Grahn, O., 1977. Macrophyte succession in Swedish lakes caused by deposition of airborne substances. Water, Air, Soil Pollut., 7: 295--305. Hargreaves, J.W., Lloyd, E.J.H. and Whitton, B.A., 1975. Chemistry and vegetation of highly acidic streams. Freshwater Biol., 5: 563--576. Harvey, H.H., 1975. Fish populations in a large group of acid stressed lakes. Verh. Int. Verein. Limnol., 19: 2406--2417. Havas, M.H. and Hutchinson, T.C., 1983. The Smoking Hills: natural acidification of an aquatic ecosystem. Nature, 301: 23--27. Hendrey, G.R. and Wright, R.F., 1976. Acid precipitation in Norway: effects on aquatic fauna. J. Great Lakes Res., 2: 192--207. Hendrey, G.R. and Vertucci, F.A., 1980. Benthic plant communities in acidic Lake Colden, New York: Sphagnum and the algal mat. In: D. Drabl#ls and A. Tollan (Editors), Ecological Impact of Acid Precipitation. SNSF report, pp. 314--315. Hitchin, G.G., Wile, I., Miller, G.E. and Yan, N.D., 1984. Macrophyte data from 46 southern Ontario soft-water lakes of varying pH. Ont. Minist. Environ. Data Report DR 84/1. Hultberg, H. and Grahn, O., 1975. Effects of acid precipitation on macrophytes in oligotrophic Swedish lakes. Proc. First Spec. Syrup. Atmospheric Contribution to the Chemistry of Lake Waters, Int. Assoc. Great Lakes Res., Sept. 28--Oct. 1, 1975, pp. 208--217. Hutchinson, G.E., 1975. A treatise in Limnology. Vol. III -- Limnologica] Botany. Wiley, New York, 660 pp. Hutchinson, T.C. and Czyrska, H., 1975. Heavy metal toxicity and synergism to floating aquatic weeds. Verh. Int. Verein. Limnol., 19: 2102--2111. Leivestad, H., Hendrey, G., Muniz, I.P. and Snekvik, E., 1976. Effects of acid precipitation on freshwater organisms. In: F.H. Braekke (Editor), Impact of Acid Precipitation on Forest and Freshwater Ecosystems in Norway. SNSF Report 6/76, pp. 87--111. Miller, G.E., Wile, I. and Hitchin, G.G., 1983. Patterns of accumulation of selected metals in members of the soft-water macrophyte flora of central Ontario lakes. Aquat. Bot., 15: 53--64. Nilssen, J.P., 1980. Acidification of a small watershed in southern Norway and some characteristics of acidic aquatic environments. Int. Rev. Gesamten Hydrobiol., 65: 177--207.
40
Okland, J. and Okland, K.A., 1980. p H level and food organisms for fish: studies of 1,000 lakes in Norway. In: D. Drabl~s and A. Tollan (Editors), Ecological Impact of Acid Precipitation. S N S F project, pp. 326--327. Oliver, B.G. and Kelso, J.R.M., 1983. A role for sediments in retarding the acidification of headwater lakes. Water, Air, Soil Pollut., 20: 379-389. Peter, R., Welsh, H. and Denny, P., 1979. The translocation of lead and copper in two submerged aquatic angiosperm species. J. Exp. Bot., 30: 339--345. Roberts, D.A., Boylen, C.W. and Singer, R., 1985. The submerged macrophyte communities of Adirondack lakes (New York, U.S.A.) of varying degrees of acidity. Aquat. Bot., 21: 219--235. Roelofs, J.G.M., 1983. Impact of acidification and eutrophication on macrophyte communities in soft waters in The Netherlands. 1. Field observations. Aquat. Bot., 17: 139--155. Roelofs, J.G.M., Schuurkes, J.A.A.R. and Smits, A.J.M., 1984. Impacts of acidification and eutrophication on macrophyte communities in soft waters. II. Experimental studies. Aquat. Bot., 18: 389--411. Singer, R., Roberts, D.A. and Boylen, C.W., 1983. The macrophyte community of an acidic lake in Adirondack (New York, U.S.A.): a new depth record for aquatic angiosperms. Aquat. Bot., 16: 49--57. Sprules, W.G., 1975. Midsummer crustacean zooplankton communities in acid-stressed lakes. J. Fish. Res. Board Canada, 32: 389--395. Stanley, R.A., 1974. Toxicity of heavy metals and salts to Eurasian watermilfoil (Myriop h y l l u m s p i c a t u m L.). Arch. Environ. Contam. Toxicol., 2: 331--341. Stokes, P.M., 1983. Response of freshwater algae to metals. Prog. Phycol. Res., 2: 87-112. Stumm, W. and Morgan, J.J., 1981. Aquatic Chemistry. Wiley, New York, 780 pp. Wile, I. and Miller, G., 1983. The macrophyte flora of 46 acidified and acid-sensitive soft water lakes in Ontario. Water Res. Branch, Ont. Minist. Environ. Rep., 35 pp. Wile, I., Miller, G.E., Hitchin, G.G. and Yan, N.D., 1985. Species composition and biomass of the macrophyte vegetation of one acidified and two acid-sensitive lakes in Ontario. Can. Field-Natural., 99(33) : in press. Wium-Andersen, S., 1971. Photosynthetic uptake of free CO 2 by roots of L o b e l i a dortmanna. Physiol. Plant., 25: 245--248. Wright, R.F., Conroy, N., Dickson, W.T., Harriman, R., Henriksen, A. and Schofield, C.L., 1980. Acidified lake districts of the world: a comparison of water chemistry of lakes in southern Norway, southern Sweden, southwestern Scotland, the Adirondack Mountains of New York, and southeastern Ontario. In: D. DrabRbs and A. Tollan (Editors), Ecological Impact of Acid Precipitation. SNSF project, pp. 377--379. Yan, N.D. and Strus, R., 1980. Crustacean zooplankton communities of acidic, metalcontaminated lakes near Sudbury, Ontario. Can. J. Fish. Aquat. Sci., 37: 2282--2293.
27
Elsevier Science Publishers B.V., A m s t e r d a m -- Printed in The N e t h e r l a n d s
RICHNESS OF AQUATIC MACROPHYTE FLORAS OF SOFT WATER LAKES OF DIFFERING pH AND TRACE METAL CONTENT IN ONTARIO, CANADA
N.D. Y A N , G.E. M I L L E R I, I. WILE s and G.G. H I T C H I N g
Ontario Ministry o f the Environment, Dorset Research Centre, P.O. Box 39, Dorset, Ont. POA lEO (Canada) ( A c c e p t e d for p u b l i c a t i o n 16 May 1985)
ABSTRACT Yan, N.D., Miller, G.E., Wile, I. and Hitchin, G.G., 1985. Richness o f aquatic m a c r o p h y t e floras o f soft w a t e r lakes of differing pH and trace m e t a l c o n t e n t in Ontario, Canada. Aquat. Bot., 23: 27--40. The richness of t h e aquatic m a c r o p h y t e floras, i.e., the total n u m b e r o f species, was assessed in 39 soft w a t e r lakes in central Ontario, Canada. T h e Cu and Ni c o n c e n t r a t i o n s and pH o f the lakes ranged f r o m 1 to 360 mg m -~, 2 to 3 7 0 0 mg m -3 and 3.9 to 7.0, respectively. T w o non-exclusive subsets of the data were e x a m i n e d to d e t e r m i n e firstly, if floral richness was related to lake pH in lakes with low Cu and Ni levels (Data Set I) and secondly, if floral richness in acidic (pH < 5.3) lakes was related to levels o f various trace metals (Data Set II). C h a r o p h y t e s were not f o u n d in lakes with pH < 5.2. In Data Set I, there was no relationship b e t w e e n the richness of t r a c h e o p h y t e s and pH, and there was a negative relationship b e t w e e n pH and b r y o p h y t e richness. Unlike p h y t o p l a n k t o n , z o o p l a n k t o n , b e n t h i c m a c r o i n v e r t e b r a t e s and fish, there was no decrease in t o t a l species richness in lakes o f pH < 5.5, as long as trace m e t a l levels were low. E x a m i n a t i o n o f Data Set II indicated t r a c h e o p h y t e richness of acidic lakes was negatively correlated w i t h Cu and Ni levels. Biological surveys of m e t a l - c o n t a m i n a t e d acidic lakes are, t h e r e f o r e , n o t of use for predicting the effects of acid deposition alone on aquatic m a c r o p h y t e s .
INTRODUCTION
Much of our information concerning the effects of acid deposition on aquatic biota has been derived from surveys of large numbers of lakes distributed along a pH gradient. Such surveys have demonstrated that the number of species of p h y t o p l a n k t o n (Almer et al., 1974; Leivestad et al., 1976), zooplankton (Sprules, 1975; Almer et al., 1978), benthic macroinvertebrates (Hendrey and Wright, 1976; Okland and q)kland, 1980) and Ontario Ministry o f the E n v i r o n m e n t , 83 A l g o n q u i n Blvd. W., T i m m i n s , Ont. P4N 2R4 Canada. Ontario Ministry of the E n v i r o n m e n t , 40 St. Clair Ave. W., T o r o n t o , Ont. M 4 V 1M2 Canada. 3 Ontario Ministry o f the E n v i r o n m e n t , P.O. Box 213, Rexdale, Ont. M9W 5L1, Canada.
0304-3770/8.5/$03.30
© 1985 Elsevier Science Publishers B.V.
28
fish (Harvey, 1975; Almer et al., 1978) is reduced when pH levels o f lake waters fall much below 6.0. While intensive surveys of small numbers o f lakes in Scandinavia have demonstrated that acidification may alter macrop h y t e assemblages (Grahn, 1977; Nilssen, 1980), the richness of the floras has rarely been assessed for a large number o f lakes. Hence, it is not yet k n o w n if the richness of macrophyte floras is positively correlated with lake pH, as is the richness of all other groups of aquatic biota. Roelofs (1983, and personal communication, 1984) indicated the richness of aquatic m a c r o p h y t e floras of lakes in The Netherlands has been reduced by acidification. However, the average pH o f these lakes is 3.9 (Roelofs et al., 1984). Lakes with pH < 4.5 are very rare in Precambrian Shield areas o f Europe and North America, where acid deposition rates are elevated (Wright et al., 1980). There has apparently been only one survey reporting the richness of macrophyte floras of r e m o t e recently-acidified lakes with a more representative range in pH. Roberts et al. (1985) surveyed 9 clear water lakes in the Adirondack Mountains of New York which ranged in pH from 4.4 to 6.9. Richness was lower in the acidified lakes, although Roberts and colleagues felt their data set was t o o small to ascribe this trend to acidity. The first purpose of this report is to determine whether macrop h y t e richness (the number of distinguishable taxa per lake) is also lower in acidified lakes in Ontario, Canada. The interpretation of results o f biological surveys of acidified lakes is frequently complicated b y the additional contamination of the acidic lakes with potentially toxic trace metals discharged from local point sources (e.g., Gorham and Gordon, 1963; Yan and Strus, 1980; Havas and Hutchinson, 1983). The second purpose of this report is to investigate whether the richness o f the m a c r o p h y t e floras of acidified lakes might be reduced b y their contamination with potentially toxic trace metals, Cu and Ni in particular, metals which unlike A1 and Mn (Wright et al., 1980) are not normally found at elevated levels in acidified lakes. METHODS
Miller et al. (1983) measured trace metal levels in tissues of aquatic macrophytes collected from 46 lakes in Ontario, ranging in pH from 3.9 to 7.0. We assembled m a c r o p h y t e richness data from 39 o f these lakes. Six of the lakes were excluded because their chemistry had been manipulated for experimental or other reasons (Hitchin et al., 1984); the seventh was excluded from our data sets (to be discussed) b y virtue o f its Ni level (> 10 mg m -3) and pH value (> 5.3). Sixteen o f the 39 lakes were located in the Sudbury area, the remaining 24, hereafter called the " D o r s e t " lakes, in south-central Ontario (Fig. 1). Each lake was visited once between May and September (usually July or August) of 1977, 1978 or 1979. Divers using snorkeling or SCUBA gear first assessed the heterogeneity of plant distribution in each lake. Consider-
29
• ,
ONTARIO
I I
,
QUEBEC
i
l
÷ O SUDBUR
°~°~< 9
oO
°
+
.:
•
0
I00 krn
._Z Fig. 1. L o c a t i o n o f s t u d y lakes. Also indicated is the z o n e within which lake p H is generally < 5.5 (cross-hatched), and d o t t e d isoline within which surface w a t e r Ni levels e x c e e d 100 mg m -3 ( f r o m Anon., 1978).
ing lake size and the results of the preliminary survey, a variable n u m b e r of sites (usually 5--15) was selected for detailed inspection. At each site, divers identified all species of submerged and floating-leaved aquatic tracheophytes, b r y o p h y t e s and charophytes found on a 10 m-wide transect extending from the shore to the maximum depth o f plant occurrence. B r y o p h y t e identifications were confirmed b y C. Manville (Royal Ontario Museum, Toronto). Macrophyte richness was not correlated with numbers of transects examined (Hitchin et al., 1984). Surface lakewater pH was measured in the field with a Radiometer Model 29 pH meter in 1--3 littoral areas, depending on lake size. The pH was measured in situ after appropriate buffer calibrations and after the readings had stabilized (< 5 min in all cases). Trace metal concentrations of surficial littoral sediments and epilimnetic or whole lake composite samples were determined as described b y Miller et al. (1983) and Hitchin et al. (1984), respectively. Divers classified each lake as highly, or n o t highly coloured (see Fig. 2).
30 TABLE
1
Data set assignment, Ni levels, pH and aquatic macrophytes
of the study lakes
x
o
~~ Data set
Ni (mg m
pH
Lake
11
3700
4.7
54*
taxa
~)
3.9
14'
300
3.9
Swal~*
*
4
4 x
II II
260
4.4
Clearwater
8
x
120
5.1
4*
7
x
I1
85
5.2
75*
9
x
l[
22
4.5
Terry
17
nd
x
x
x
Freela
17
x
x
Bell
32
x
x
x
x x
x x x
x
x
x
x x x
x
4.8
George
23
x
x
x
II
6
4.6
Carlyle
30
x
x
x
II II II II
4 ~2 -3 -:2
4.0 4.2 5.2 5.3
Krarner * * Horn" plastic Cinder
5 17 17 15
x
11
,:2
5.3
Crosson
19
x
x
II
<2
4.4
Axe
24
x
x
x
11
-2
5.3
Fawn
24
x
x
x
x x x
5.5
Leonard
16
x
x
x
,4
5.5
Heney
28
x
x
x
x
<~3
5.6
MeKay
27
x
x
x
42
5.6
Healey*
22
x
<3 <2 <~2 2
5.7 5.7 5.8 5.9
Leech 94* Gud feather ÷ Moot*
17 22 19 15
2
5.0
x x
x x
x x
x
x x
x x
x x x
x
x x
x x
x x x
x
x x
x x x
x
x x x
x x
x
x
x x x
x x
x
x
x
x x x
x
x
x x x x
x x x x
x
x
x
x x x
x
x x x
x
x
x
x x
Hillman +
19
x
x
6.4
Clear
l 5
x
x
x
6.4 6.4 6.5
Brandy Otter Little Otter R e d Chalk Frond Harp LittLe C l e a r Solitaire
18 27 25 12
x
x x x
x x x
44
x
x
x x
codes
follow
**Local name only. ***Ni ieveB una~ilable eoloured
Gorham but
lakes.
assumed
and
Gordon tt) b~
x
x
x
x
x
x
x
x x x
x
x
x x
x
x
x
x
x
x
x
x
x
x x
x
<']0
x
(I 9 6 3 } , mR
m
' based
on
hike
location.
x
x x x
x
x
x
x
x x
x
x x
x x x x x x
x x
x
x
x
x
x
x x
x
x
x x x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x x x
x
x
x
x
x
x
x
x x
x
x
x
x
x
x x
x
x x
x
x x
x
x
x
x x
x
x x x
x
x x x
x
x
x x x
x
x X
x
x x x
x
x
x
x x
x x
x
x x
x x
x
x x
x
x x
x
x
x
x
x x
x
x x x
x
x
x x
x x
x
x
x x
x x
x
x
x x
x
x x
x x x
x x x x x
x
6.3
x
x x
x
x x
x
x x
x x
x x
x
x x
X X
x x
x
I1 16 6
x
x
18
6.8 6.9 7.0
x x
19
<2 "~2 < 2
x
x x x x
2]
x
x x
103"
x
x x
x x
Dickie
**
x
x x
Chub
6.6 6.6
x
x
x x
x x x
x x
6.0
--*
* Highly
x
x
x x x
x x x
x
x
x
x
x x
x x
x x
x
x x x
6.2
2 --** * --*** 2
o
x
x x x
x x
3
< 2
x x
x
x x
• 2
--***
*Numeric
x
x ×
x x
fi
<
o
x x
II
3
o
x
x
x
4.8
x
x x
x
4.3
+
o
x
660
1l
0
x
x
[[ II
• :20
0
~
3
÷
o
x
x
x x x
x
x
~ x x
x
~x x
x
x
x
x
~t
xx~
Sull.
P. B e a u v . a m b l y p h y l l u r a
Spha~um
tere~ ( S c h i r a p . ) ~ t n g s t r .
Sphagnum sub~ecundum N c e s & H o r n s c h .
Sphagnum r e c u r u u m
Sphagnum pylae|ii Brld.
Sphagnum palu#tre L.
Sphagnum mqJua ( R u s s . ) C, Jcns.
Sphagnum ftmbr~atum W i l s .
Sphagnum cuJpidatum E h r h .
(Row.)Warnst.
Pohlia nu tanm ( H e d w .) L i n d b . s c h i m p e r i i ( H e d w .) L i n d b .
Hygroomblymtegium tenox ( H e d w . ) Jenn.
Gymnocolea inflata ( H u d s . ) D u m .
Fontinal~ nova-angliae
Fontinalis durlaei S c h i m p .
Fonttnalil hypnotde# R . H a r t m a n n
FontinoJim anttpyrettca H e d w .
Fis#iden# fort tanuJ ( B . P y l . ) S t e u d .
Drepanocladus sp.
CIadopodiella nultanw (Nees) Buch.
Nitella sp,
d u r i a e i F.
Bruz.) Ag.
Nite~ia tenuluima ( D e s v . ) K / / t ; ' .
Nitella gracili~ (Sin,) A g ,
Nitella furcata ( R o x h . e x
Nitella flexilis ( L . ) A g .
Chara sp.
ValllzneMa ame~cana M i c h x
UO'icular4a uutgarim L.
Utr~cularia re#uplnata B . D . G r e e n
Utricularta purpurea W a l t ,
Utricu~aria m i n o r L.
Utrtcular~a intermedi= H a y n e
Utr~culari* gtbba L.
Utrtcularla c o ~ u t a M i c h x
Sparganium sp.
Sagittar~a sp.
Ranuneulu~ ~ep Fans L.
Potamogeton uo,w#eyl R o b b i n s
Potamogeton #pirillbl T u c k e r m .
Potamogeton robbln#ff O a k e s
Potomageton Hchard#onil ( B e r n . ) Rydb.
Potaenogeton obtu,ifoltus M e r t . & K o c h . PotamoEeton puMIlus L.
O0 b.a
32
Description o f the study lakes The pH and Ni levels o f the study lakes are listed in Table I. Lake pH values ranged from 3.9 to 6.6 among the Sudbury lakes and from 4.2 to 7.0 among the Dorset lakes. Several o f the acidic Dorset lakes had coloured waters, indicating their acidity was partially attributable to organic acids. Levels of Ca ranged from 2 to 5 mg 1-1 among the Dorset lakes and from 2 to 15 mg 1-1 among the Sudbury lakes. The elevated levels were f o u n d in the acidic lakes near Sudbury and were attributable to elevated rates o f weathering of watershed materials produced by input of acids (Dillon, 1983). Levels of Cu and Ni were greatly elevated in lakes within 25 km o f the Sudbury smelting complex (Fig. 1 and Conroy et al., 1975). Levels were much lower in the acidic lakes located further from the city and in all lakes with pH values > 5.5. Eriocaulon septangulare With. and Eleocharis acicularis (L.) R. and S. were the most frequently encountered species, occurring in over 75% of the lakes. An additional 14 species, occurred in half or more t h a n half o f the lakes (Table I). Roberts et al. (1985) identified Potamogeton confervoides Reichenb. and Sphagnum spp. as characteristic o f acidic lakes in the Adirondacks. This agrees with their occurrence in the Ontario study lakes. P. confervoides was f o u n d in 9 lakes with a median pH of 4.8 (range 4.3--5.6). Sphagnum spp. were f o u n d in 15 lakes ranging in pH from 4.2 to 6.4 with a median pH of 5.3. While various Utricularia species were observed in the study lakes, Utricularia geminiscapa Benj. was not observed. This species is characteristic o f acidic lakes in the Adirondacks (Roberts et al., 1985). The plant assemblages of the study lakes are described in detail elsewhere (Wile and Miller, 1983). RESULTS
The survey data were organized into 2 subsets to address the 2 stated purposes. To examine the influence o f depressed pH itself on m a c r o p h y t e richness, data from those 30 lakes with lake water Ni levels < 10 mg m -3 were examined (Data Set 1). As Ni levels in lake water were positively correlated (P < 0.01) with levels of Ni (r = 0.95) and Cu (r = 0.91) in littoral sediments and with Cu levels in lake water (r -- 0.98, n = 40), the Ni levels in lake water were assumed to provide a satisfactory index o f Ni and Cu contamination of waters b o t h above and below the sediment surface. Copper can be extracted both from the water and the sediments by macrophytes (Peter et al., 1979). The 18 lakes with pH ~< 5.3 were operationally termed " a c i d i c " and formed Data Set 2. This set o f data was assembled to examine the relationship between m a c r o p h y t e richness in acidic lakes and habitat metal contamination. A separatory pH of 5.3 was chosen for 2 reasons. It approxi-
33
mates the equivalence point o f alkalinity titrations of soft waters; hence, soft water lakes with pH ~< 5.3 should have measurable free acidity ( S t u m m and Morgan, 1981). Secondly, changes in macrophyte c o m m u n i t y structure have been observed at approximately this pH (Roberts et al., 1985). By design, ranges o f Cu and Ni levels in lake water were low in Data Set 1 (1--3 mg m -3 and 2--6 mg m -3, respectively), b u t ranges of levels o f o t h e r trace metals were quite large (3--210 mg Zn m -3, 18--135 mg Mn m -3, 30--1150 mg Fe m -3 and 10--360 mg A1 m-3). The lakes in Data Set 1 ranged in pH from 4.0 to 7.0, and supported an average of 20.1 (range 5--44) species of aquatic macrophytes, o f which 16.9 (4--39) were tracheophytes, 2.4 (0--6) were b r y o p h y t e s and 0.8 (0--3) were charophytes. Significantly more species of aquatic b r y o p h y t e s (Musci) were found in the acidic lakes (Table II, Fig. 2). A positive correlation between b r y o p h y t e richness and lake area was also observed. This was n o t an indirect consequence o f a lake area--lake pH relationship (r = 0.26, P 0.10). It m a y simply reflect increased habitat area in larger lakes. Nevertheless, a greater relative and absolute richness of species o f the k n o w n acidophilous taxon, Sphagnum, in lakes of pH < 5.0 (Table I and Wile and Miller, 1983) suggests that it is probably lake acidity that directly or indirectly controls b r y o p h y t e richness. TABLE II Correlation coefficients between aquatic plant richness and selected limnological variables for lakes in Data Set 1. Significance at 95 and 99% indicated by single and double asterisks, respectively Parameter
Bryophyte
pH --0.47** A1 --0.05 Ca --0.21 Conductivity --0.11 Lake area 0.38*
Charophyte
Tracheophyte
Macrophyte
0.38* --0.11 0.50** 0.63" * 0.03
0.01 --0.28 0.60** 0.31 0.21
--0.04 --0.24 0.55** 0.32 0.25
Charophytes were observed in 15 lakes, all with pH > 5.2 (Fig. 2). In his review, Hutchinson (1975) noted that the majority of charophyte species are found in alkaline waters. The positive correlations between c h a r o p h y t e richness and pH, Ca and conductivity (Table II) are consistent with this observation. The richness of aquatic tracheophytes was not significantly correlated with pH, conductivity or lake area in Data Set 1 (Table II). Additionally there was not a significant correlation with A1, the toxic metal p r o b a b l y directly responsible for lowering the richness of fish communities in acid lakes (Baker and Schofield, 1982). Richness was correlated with Ca levels. This was in part attributable to the unusually high richness o f F r o o d Lake
34 L~ < w I DO n~ •
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pH Fig. 2. Scattergrarns of pH vs. numbers o f species o f aquatic Tracheophyta, Musci and Charophyceae in each lake in Data Set 1. Highly coloured lakes are indicated by open circles for the Tracheophyta. Numbers indicate the number o f coincident points.
(39 taxa), the lake with the highest Ca levels. However, elimination of Frood Lake from the data matrix did not eliminate the correlation o f tracheophyte richness with Ca (r = 0.40, P < 0.05). There were 9 coloured lakes in Data Set 1. The pH o f these lakes ranged from 4.2 to 6.3. The tracheophyte richness o f the coloured lakes was similar to that of the lakes with greater transparency (Fig. 2), indicating that tracheophyte richness in recently acidified, clear water lakes is apparently not different from acidic lakes with coloured waters. Such lakes presumably have been acidic for longer periods o f time. There may, o f course, be differences in species assemblages between the coloured and clear water acidic
35 lakes; however, as such differences were not detected in the ordination of species presence--absence data {Wile and Miller, 1983), any differences must be small. Because of the opposing patterns of bryophyte and charophyte richness and because of the poor correlation between tracheophyte richness and 30-
30-
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Fig. 3. Scattergrams of tracheophyte richness vs. water column levels of Fe, Zn, AI, Mn, C u a n d Ni for lakes in Data Set 2. Three values were unavailable for the Al plot.
36 pH, there was no relationship between total m a c r o p h y t e richness and pH (r = --0.04, P > 0.10). The analysis of Data Set 1 produced significant correlations b e t w e e n pH and the richness of b r y o p h y t e and charophyte, but n o t tracheophyte floras. Therefore, we chose to examine only t r a c h e o p h y t e data in Data Set 2, the set o f all lakes with pH ~< 5.3. Lake pH varied from 4.0 to 5.3 for the 18 lakes in Data Set 2 (Table I) and was n o t significantly correlated (r = 0.35, P > 0.10) with tracheophyte richness. Ranges o f Zn, Mn, Fe and A1 were similar in this data set to those in Data Set 1 (3--210, 43--330, 30--1150 and 27--360 mg m -3, respectively). Ranges of Cu (1--360 mg m -3) and Ni (2--3700 mg m-3), were, of course, much wider. Scattergrams o f tracheophyte richness and lakewater levels o f Fe, Zn, Mn, A1, Cu and Ni are presented in Fig. 3. No pattern is evident for Fe and Zn. For A1 and Mn, there is some indication o f reduced richness at the highest metal levels. There is a clear pattern o f decreasing richness with increasing levels of Cu and Ni. DISCUSSION
Acidification and macrophyte richness Summarizing the results of various surveys o f naturally acidic lakes, Hutchinson (1975) indicated that charophytes are generally not observed in lakes of pH < 5.5. Therefore, the absence of charophytes from lakes of pH < 5.2 in this survey (Fig. 2) is consistent with k n o w n patterns of occurrence of the group. As charophytes similarly have n o t appeared on any m a c r o p h y t e species lists from aquatic plant surveys of recently acidified lakes (Hendry and Vertucci, 1980; Eriksson et al., 1983; Roelofs, 1983; R o b e r t s et al., 1985), it may be postulated that the charophycean flora of soft water lakes will probably be eliminated should lake pH fall to 5.0 or below. With the exception of the charophytes, which made only a small contribution to total macrophyte richness, there was no evidence of reduced m a c r o p h y t e richness in lakes of low pH (Fig. 2). Therefore, in terms o f richness, m a c r o p h y t e s differ from plankton, benthic macroinvertebrates and fish, all o f which have reduced richness in acidic lakes (Dillon et al., 1984). Most of the m a c r o p h y t e species found in these s t u d y lakes actually occurred over the entire pH range of 4.0--7.0 (Table I). While this seems to indicate that most tracheophyte species observed in soft water Shield lakes are tolerant of substantial pH depressions, their tolerance may be more apparent than actual. The roots o f aquatic m a c r o p h y t e s of soft water lakes are the primary site of uptake and exchange of nutrients and major ions (Wium-Andersen, 1971; Carignan and Kalff, 1980). Below the t o p few centimetres, the pH o f the porewater of littoral zone sediments in acidic lakes is generally much higher than that o f the overlying water (Hultberg
37
and Grahn, 1975; Roberts et al., 1985). Hence, most exchange o f materials in these m a c r o p h y t e s probably takes place across membranes which are n o t exposed to strongly acidic water. Collins et al. (1981) similarly indicated that the isolation of benthic invertebrates found beneath the surface o f the sediments o f lakes (infauna) from water of lowest pH might explain w h y these organisms are n o t as severely affected b y the acidification o f lakes as epifaunal invertebrates. It is n o t surprising that acidification has reduced the richness o f lakes in The Netherlands, b u t not in Ontario. With an average pH of 3.9, those lakes studied b y Roelofs et al. {1984) are more acidic than the lakes studied herein. Further, the Dutch lakes have p r o b a b l y been acidic for longer periods of time (Beamish and Harvey, 1972; Roelofs, 1983). As acidification of sediment pore water lags behind that of lake waters (Oliver and Kelso, 1983) and as acidification is probably a more recent p h e n o m e n o n in Ontario, it may be that detrimental impacts on the m a c r o p h y t e floras of Ontario lakes are yet to be observed. Roberts et al. (1985) surveyed 9 lakes in the Adirondack Mountains of New York which ranged in pH from 4.4 to 6.9 and noted that total macrophyte richness was lower in the acidic than in the non-acidic lakes. In contrast, results of our survey indicate that the depressions in pH of the magnitude o f that usually reported in Canadian Shield Lakes (pH > 4.5, Wright et al., 1980) probably will n o t result in reductions in aquatic b r y o p h y t e or tracheophyte richness. Intensive m a c r o p h y t e surveys have indicated that acidification m a y alter the biomass (Wile et al., 1985) and composition of m a c r o p h y t e communities (Hultberg and Grahn, 1975; Roelofs, 1983) and the distribution of macrop h y t e species with depth (Singer et al., 1983). However, for lakes in Ontario, it does n o t appear that acidification of lakes has reduced macrophyte richness.
Trace metal contamination o f acidic lakes and macrophyte richness There was no relationship between Fe and Zn levels and tracheophyte richness among the acidic lakes (Fig. 3). Relationships were not anticipated with Fe as it is not a particularly p h y t o t o x i c element (Stokes, 1983) and as the observed Fe concentrations were n o t unusually high for Shield lakes. Concentrations of Zn were much higher in some of the acidic lakes in Data Set 2 than typically observed in either acidic or non-acidic Shield lakes (Borg, 1983). Nevertheless, toxic effects of Zn on macrophytes are not evident until levels are orders of magnitude higher than Zn levels recorded in this study (Stanley, 1974). Hence, a relationship between Zn levels and t r a c h e o p h y t e richness was not anticipated. Tracheophyte richness was low in lakes with high levels of Cu and Ni, supporting Gotham and Gordon's (1963) hypothesis that their observed positive relationship between m a c r o p h y t e richness and distance of lakes
38
from Sudbury might be related to levels of Cu and Ni. Tracheophyte richness was also reduced in a few lakes with approximately 300 mg m -3 or more o f A1 and 150 mg m -3 or more o f Mn (Fig. 3). It is, o f course, not possible to use these survey data to identify the cause of the unusually low tracheop h y t e richness in the metal-contaminated acidic lakes. However, if metal contamination is the cause, there are three reasons w h y Cu and Ni are more likely causative agents than A1 or Mn. Firstly, Scandinavian acidic lakes o f t e n have levels of A1 and Mn similar to those observed in the acidic lakes in Data Set 2 (Almer et al., 1978; Wright et al., 1980; Borg, 1983), b u t t h e y do not have elevated levels o f Cu and Ni, nor do they have unusually low m a c r o p h y t e richness (Eriksson et al., 1983). Secondly, Cu is much more p h y t o t o x i c than A1 and Mn (Stokes, 1983; Stanley, 1974) and Ni can seriously damage aquatic macrophytes at levels observed in Data Set 2, a b o u t 300 mg m -3 (Hutchinson and Czyrska, 1975; Brown and Rattigan, 1979). Finally, as Hutchinson and Czyrska (1975) reported synergism in the toxicity of Cu and Ni to aquatic plants, the presence o f high concentrations o f b o t h metals in several lakes in Data Set 2 probably exacerbates their toxic effects on macrophytes. Biological surveys o f acidic freshwaters have o f t e n included lakes or streams which are both acidic and contaminated with trace metals attributable to local industrial operations (Yan and Strus, 1980) to natural fumigations with metal-laden particulates (Havas and Hutchinson, 1983) or to drainage from acidic mine tailings (Hargreaves et al., 1975). This study indicates that the results of such surveys m a y n o t be o f use in predicting the consequences for aquatic biota of excessive inputs of strong acids alone. ACKNOWLEDGEMENTS
Appreciation is extended to W. Kerr, J. Scholer, S. Painter and D. Snell for assistance in the field and laboratory, to J. Roelofs for advice on use of his data and to B. Barley, R. Hall, D. R o b e r t s and P. Stokes for fruitful discussions and reviews of early drafts o f the manuscript.
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39 Borg, H., 1983. Trace metals in Swedish natural freshwaters. Hydrobiologia, 101: 27--34. Brown, B.T. and Rattigan, B.M., 1979. Toxicity of soluble copper and other metal ions to Elodea canadensis. Environ. Pollut., 20: 303--314. Carignan, R. and Kalff, J., 1980. Phosphorus sources for aquatic weeds: water or sediments? Science, 207: 987--989. Collins, N.C., Zimmerman, A.P. and Knoechel, R., 1981. Comparisons of benthic infauna and epifauna biomasses in acidified and nonacidified Ontario lakes. In: R. Singer (Editor), Effects of Acid Precipitation on Benthos. North Am. Benthol. Soc., Hamilton, New York, pp. 35--48. Conroy, N., Hawley, K., Keller, W. and Lafrance, C., 1975. The influence of the atmosphere on lakes in the Sudbury area. Proc. First Spec. Symp. Atmospheric Contribution to the Chemistry of Lake Waters, Int. Assoc. Great Lakes Res., Sep. 28--Oct. 1, 1975, pp. 146--165. Dillon, P.J., 1983. Chemical alterations of surface waters by acidic deposition in Canada. Water Qual. Bull., 8: 127--132. Dillon, P.J., Yan, N.D. and Harvey, H.H., 1984. Acidic deposition: effects on aquatic ecosystems. CRC Crit. Rev. Environ. Control, 13: 167--194. Eriksson, F., H6mstrSm, E., Mossberg, P. and Nyberg, P., 1983. Ecological effects of lime treatment of acidified lakes and rivers in Sweden. Hydrobiologia, 101: 145--164. Gorham, E. and Gordon, A.G., 1963. Some effects of smelter pollution u p o n aquatic vegetation near Sudbury, Ontario. Can. J. Bot., 41: 371--378. Grahn, O., 1977. Macrophyte succession in Swedish lakes caused by deposition of airborne substances. Water, Air, Soil Pollut., 7: 295--305. Hargreaves, J.W., Lloyd, E.J.H. and Whitton, B.A., 1975. Chemistry and vegetation of highly acidic streams. Freshwater Biol., 5: 563--576. Harvey, H.H., 1975. Fish populations in a large group of acid stressed lakes. Verh. Int. Verein. Limnol., 19: 2406--2417. Havas, M.H. and Hutchinson, T.C., 1983. The Smoking Hills: natural acidification of an aquatic ecosystem. Nature, 301: 23--27. Hendrey, G.R. and Wright, R.F., 1976. Acid precipitation in Norway: effects on aquatic fauna. J. Great Lakes Res., 2: 192--207. Hendrey, G.R. and Vertucci, F.A., 1980. Benthic plant communities in acidic Lake Colden, New York: Sphagnum and the algal mat. In: D. Drabl#ls and A. Tollan (Editors), Ecological Impact of Acid Precipitation. SNSF report, pp. 314--315. Hitchin, G.G., Wile, I., Miller, G.E. and Yan, N.D., 1984. Macrophyte data from 46 southern Ontario soft-water lakes of varying pH. Ont. Minist. Environ. Data Report DR 84/1. Hultberg, H. and Grahn, O., 1975. Effects of acid precipitation on macrophytes in oligotrophic Swedish lakes. Proc. First Spec. Syrup. Atmospheric Contribution to the Chemistry of Lake Waters, Int. Assoc. Great Lakes Res., Sept. 28--Oct. 1, 1975, pp. 208--217. Hutchinson, G.E., 1975. A treatise in Limnology. Vol. III -- Limnologica] Botany. Wiley, New York, 660 pp. Hutchinson, T.C. and Czyrska, H., 1975. Heavy metal toxicity and synergism to floating aquatic weeds. Verh. Int. Verein. Limnol., 19: 2102--2111. Leivestad, H., Hendrey, G., Muniz, I.P. and Snekvik, E., 1976. Effects of acid precipitation on freshwater organisms. In: F.H. Braekke (Editor), Impact of Acid Precipitation on Forest and Freshwater Ecosystems in Norway. SNSF Report 6/76, pp. 87--111. Miller, G.E., Wile, I. and Hitchin, G.G., 1983. Patterns of accumulation of selected metals in members of the soft-water macrophyte flora of central Ontario lakes. Aquat. Bot., 15: 53--64. Nilssen, J.P., 1980. Acidification of a small watershed in southern Norway and some characteristics of acidic aquatic environments. Int. Rev. Gesamten Hydrobiol., 65: 177--207.
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Okland, J. and Okland, K.A., 1980. p H level and food organisms for fish: studies of 1,000 lakes in Norway. In: D. Drabl~s and A. Tollan (Editors), Ecological Impact of Acid Precipitation. S N S F project, pp. 326--327. Oliver, B.G. and Kelso, J.R.M., 1983. A role for sediments in retarding the acidification of headwater lakes. Water, Air, Soil Pollut., 20: 379-389. Peter, R., Welsh, H. and Denny, P., 1979. The translocation of lead and copper in two submerged aquatic angiosperm species. J. Exp. Bot., 30: 339--345. Roberts, D.A., Boylen, C.W. and Singer, R., 1985. The submerged macrophyte communities of Adirondack lakes (New York, U.S.A.) of varying degrees of acidity. Aquat. Bot., 21: 219--235. Roelofs, J.G.M., 1983. Impact of acidification and eutrophication on macrophyte communities in soft waters in The Netherlands. 1. Field observations. Aquat. Bot., 17: 139--155. Roelofs, J.G.M., Schuurkes, J.A.A.R. and Smits, A.J.M., 1984. Impacts of acidification and eutrophication on macrophyte communities in soft waters. II. Experimental studies. Aquat. Bot., 18: 389--411. Singer, R., Roberts, D.A. and Boylen, C.W., 1983. The macrophyte community of an acidic lake in Adirondack (New York, U.S.A.): a new depth record for aquatic angiosperms. Aquat. Bot., 16: 49--57. Sprules, W.G., 1975. Midsummer crustacean zooplankton communities in acid-stressed lakes. J. Fish. Res. Board Canada, 32: 389--395. Stanley, R.A., 1974. Toxicity of heavy metals and salts to Eurasian watermilfoil (Myriop h y l l u m s p i c a t u m L.). Arch. Environ. Contam. Toxicol., 2: 331--341. Stokes, P.M., 1983. Response of freshwater algae to metals. Prog. Phycol. Res., 2: 87-112. Stumm, W. and Morgan, J.J., 1981. Aquatic Chemistry. Wiley, New York, 780 pp. Wile, I. and Miller, G., 1983. The macrophyte flora of 46 acidified and acid-sensitive soft water lakes in Ontario. Water Res. Branch, Ont. Minist. Environ. Rep., 35 pp. Wile, I., Miller, G.E., Hitchin, G.G. and Yan, N.D., 1985. Species composition and biomass of the macrophyte vegetation of one acidified and two acid-sensitive lakes in Ontario. Can. Field-Natural., 99(33) : in press. Wium-Andersen, S., 1971. Photosynthetic uptake of free CO 2 by roots of L o b e l i a dortmanna. Physiol. Plant., 25: 245--248. Wright, R.F., Conroy, N., Dickson, W.T., Harriman, R., Henriksen, A. and Schofield, C.L., 1980. Acidified lake districts of the world: a comparison of water chemistry of lakes in southern Norway, southern Sweden, southwestern Scotland, the Adirondack Mountains of New York, and southeastern Ontario. In: D. DrabRbs and A. Tollan (Editors), Ecological Impact of Acid Precipitation. SNSF project, pp. 377--379. Yan, N.D. and Strus, R., 1980. Crustacean zooplankton communities of acidic, metalcontaminated lakes near Sudbury, Ontario. Can. J. Fish. Aquat. Sci., 37: 2282--2293.