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Patterns of vegetation change in Lake Wingra following a Myriophyllum spicatum decline
Aquatic Botany, 46 (1993) 325-340
325
0304-3770/93/$06.00 © 1993 - Elsevier Science Publishers B.V. All rights reserved
Patterns of vegetation change in Lake Wingra following a Myriophyllum spicatum decline Anett S. Trebitz *,a, Stanley A. Nichols b, Stephen R. Carpenter~, Richard C. Lathrop c "Centerfor Limnology, Universityof Wisconsin, 680 N. Park Sreet, Madison, W15 3 706, USA bWisconsin Geologicaland Natural History Survey, 3817 Mineral Point Road, Madison, W153705, USA CBureauof Research, WisconsinDepartment of Natural Resources, 3911 Fish Hatchery Road, Madison, W153711, USA (Accepted 15 June 1993)
AbsUa~
The invading aquatic plant, Myriophyllum spicatum L. is a management concern in many North American lakes because it replaces native species and because its dense growth can be a nuisance to lake users. It is common for M. spicatum to expand quickly upon reaching a lake, remain the most abundant littoral plant for a number of years, and then decline rather rapidly. This pattern held true for Lake Wingra, Dane County, WI, where M. spicatum dominated the littoral vegetation during the late 1960s, but abruptly declined during the 1970s. In this paper, we explore the changes in the Lake Wingra plant community that have occurred in the wake of the M. spicatum decline. We present results of 1991 and 1992 vegetation surveys indicating that M. spicatum, while no longer the dominant macrophyte, remains an important member of the Lake Wingra plant community. It and Ceratophyllum demersum L. now make up roughly equal parts of the littoral vegetation, and native species, rare or absent during the 1960s, are growing well. By comparing current plant distributions with those found earlier, we examined probable causes for the M. spicatum decline; no single factor seemed to be responsible. The miifoil decline in Lake Wingra has been sustained over roughly two decades while the native vegetation has expanded.
Introduction Species invasions and their effects are of considerable interest to ecologists (Elton, 1958; Johnstone, 1986; Drake et al., 1989; Lodge, 1993; and references therein). Often, the expansion of a disruptive invader is closely monitored, but once the invasion 'problem' diminishes, much less attention is given to the system's recovery. This pattern has-applied to the invasion of Eurasian milfoil (Myriophyllum spicatum L. ) in many North American lakes. It has been observed repeatedly that M. spicatum exhibits 'boom and bust' *Corresponding author.
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growth patterns, but it is not known what precipitates a decline. Proposed agents include nutrient depletion, light levels, competition, herbivory, parasites or pathogens, and management effects (Carpenter, 1980; Painter and McCabe, 1988; Smith and Barko, 1990). Often, competing factors cannot be separated because of insufficient information on lake conditions prior to the arrival of milfoil. There is also generally little information on plant community changes following a milfoil decline (Smith and Barko, 1990 ). In Lake Wingra, WI, we are fortunate to have records of macrophyte composition during the height of milfoil growth and for the first years of its decline ( 1969-1977 ). In this paper, we compare the results of 1991-1992 vegetation surveys with earlier data to describe how the vegetation has changed, approximately 15 years after milfoil began to decline. Probable causes of the decline are examined by comparing the present distribution of M. spicatum with patterns that might be expected from various causative agents. Lake history Lake Wingra is the smallest of several lakes within the Madison metropolitan area of Dane County, WI. This alkaline, eutrophic lake has a maximum depth of 4.3 m, a mean depth of 2.7 m, and a surface area of just under 140 ha (Baumann et al., 1974). About a third of its area is littoral (Adams and Prentki, 1982). Its shoreline is almost entirely within parks, while much of its watershed is residential. A few houses line its northern shore, as do a public swimming beach and a boat launch. Historically, Lake Wingra had a diverse plant community, with many floating and emergent species as well as submerged macrophytes. Prior to the 1900s, the area was marshy and the lake shore was extensively bordered by emergent plants (Baumann et al., 1974). Wild celery ( Vallisneria americana Michx. ) was reportedly quite abundant. Dredging and filling in the early 1900s and lowered water levels eliminated much of the marsh on the northeast and east sides of the lake. The carp (Cyprinus carpio L. ), abundant by 1915, also had a negative influence on vegetation (Baumann et al., 1974). Macrophyte abundances and diversities generally declined over the first half of the century, although Najas flexilis Rostk. & Schmidt, Potamogeton pectinatus L., Potamogeton natans L. and Ceratophyllum demersum L. were still common in 1943 (Zimmermann, 1953 ). V. americana was lost from the lake during this period. In the 1950s, carp control measures were taken and the macrophyte community recovered somewhat. At this time, the vegetation was dominated by pondweeds (Potamogeton spp. ), the native milfoil Myriophyllum sibiricum Kom. (formerly called Myriophyllum exalbescens Fernald), and coontail ( C. demersum) (Zimmermann, 1953 ). It is not known exactly when M. spicatum invaded Lake Wingra, but it must have expanded rapidly, since it was not reported prior to the 1960s, yet was
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abundant by 1966 (Nichols and Mori, 1971 ). A survey of the lake flora in 1969 documented the loss of M. sibiricum, V. americana, and several Potamogeton species along with the dominance ofM. spicatum (Nichols and Mori, 1971 ). Eurasian milfoil was also a problem in other Madison lakes during this time (Lind and Cottam, 1969 ) and resulted in a dramatic increase in the biomass of weeds (principally milfoil) cut and removed from the lakes (Lathrop, 1989 ). By the mid 1970s, milfoil had declined dramatically in all the Madison lakes for reasons that remain unclear (Carpenter, 1980; Lathrop, 1989). Methods
1991 and 1992 vegetation surveys We sampled Lake Wingra vegetation using a rapid survey method, whereby a weighted, double-headed rake is thrown into the water from a boat, dragged about 2 m across the bottom, and then pulled up by means of an attached line (Deppe and Lathrop, 1993 ). Plants recovered from the rake tines are identified to species and assigned an abundance ranging from 0 (not present) to 5 (very abundant). Abundances are based roughly on the biomass of plants of each species present in the sample. Sampling was done along transects perpendicular to the shore, with four replicate samples taken at each 0.5 m depth across the littoral zone. While this is not a method that yields precise abundance data, as is the case for example with sampling quadrats or transects by SCUBA, the rake method has the advantage of being much faster, and of covering more area, thereby increasing the chances of encountering rare species. Our 1991-1992 rake transects covered historically studied areas (Nichols and Mori, 1971 ) and avoided the swimming beach and boat launch along the northern shoreline (Fig. 1 ). The results of the two 1991 surveys show that the rake survey method is robust and repeatable: lakewide relative frequencies for the common species are quite similar (Table 1 ), and depth distributions of the dominant species in 1991 are comparable (Fig. 2). The 1991 surveys are in less agreement on the abundances of rare species, in part because our 16 August survey crew had less experience in identifying the rarer species. Also, rare species are usually more variable among surveys. We used our 6 August 1991 data for a comparison with frequencies from earlier years, but pooled the two 1991 surveys for distribution analyses of the common species. Limited biomass sampling was done in the west end of the lake concurrent with the 6 August 1991 rake survey. The total dry weight of all plants removed from 0.1 m 2 quadrats was recorded (plants dried for 48 h at 60°C). A total of 20 quadrats were sampled, ten taken randomly along each of the 1.5 and 2.5 m depth contours in the west end of the lake.
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approximatelocationofsurveytransects ...................................... 6 A u g 9 1
IN ~
Fig. 1. Map of Lake Wingra showing location of survey transects. Depth contours (m) are approximate. Table 1 Relative frequency of macrophytes found by two 1991 rake surveys. The 6 August 1991 survey also found Potamogeton natans L., Potamogeton zosteriformis Fernald, Chara spp., Myriophyllum sibiricum Kom., Potamogeton illinoensis Morong, Vallisneria americana Michx. and Utricularia macrorhiza LeConte Species
Myriophyllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Ceratophyllum demersum L. Elodea canadensis Michx. Heteranthera dubia (Jacq.) MacM. Potamogeton crispus L. Najasflexilis Rostk. & Schmidt Potamogeton foliosus Raf. Potamogeton richardsonii Rydb.
Per cent relative frequency 6 Aug.
16 Aug.
25.9 6.2 5.1 26.6 2.3 7.8 3.1 5.3 2.6 0.7
24.0 5.8 7.9 27.1 2.7 9.1 2.1 10.0 2.1 5.2
Comparison with earlier data Lakewide relative frequency data were used to compare the plant community found in this study with that measured in 1969 (Nichols and Mori, 1971 ), 1970 (Nichols, 1971 ), and 1977 (Kimbel and Johnson, cited in Carpenter,
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A.S. Trebitz et aL /Aquatic Botany 46 (1993) 325-340 6 Aug 91 © © 16 Aug 91 -. ...........• 18 Aug 92
,,,,"
',.
.~_
~2 V
8
H. dubia
g2 IV. flex#is .5
A
~
.
P. pectinatus
oT--
o.s
=
1.o
1.s 2.o 2.s water depth (m)
.
3.0
Fig. 2. The depth distribution of dominant plants (more than 5% relative abundance) in 1991 and 1992.
1980). Relative frequency data came from line-intercept surveys in 1969 and 1977, from quadrats in 1970, and from our rake surveys in 1991 and 1992. Since relative frequencies are computed from presence/absence information, comparison of these data is unaffected by differences among how the various methods measure abundance. The complete plant data set for 1977 has been lost, but species cited in Carpenter (1980) account for 97.1% of the relative frequency. Detailed measurements of lakewide milfoil biomass and sediment organic content and texture (percentage of particles less than 0.5 mm in diameter) were made in Lake Wingra in 1970 at the height of the milfoil invasion (Nichols, 1971 ). To compare 1970 milfoil biomasses with our own limited 1991 biomass samples, we computed 1970 biomass means and standard errors using only samples from approximately the 1.5 and 2.5 m depths in the west end of the lake. We also analyzed vegetation changes through correlation and
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map analyses that used spatial information in the 1970 and 1991-1992 data. We divided the lake into eight broad regions, chosen through a combination of the 1991-1992 transect coverage and the littoral zone topography (Fig. 1 ). Within each region, six 0.5 m depth classes were defined, resulting in a total of 48 map segments around the lake. Plant data for 1991-1992, and 1970 plant and sediment data were then assigned to these segments based on depth and position along the shoreline. Correlation analyses and distribution maps are based on the mean plant or sediment value within each segment. We also compared our 1991-1992 plant data with macrophyte community distributions established in 1969 by Nichols and Mori ( 1971 ). They identified five macrophyte communities in Lake Wingra: the shallow and deepwater milfoil communities, the Nuphar community, the Nymphaea community, and a mixed community. We adapted their classification to apply to conditions currently found in the lake. Because coontail (C. demersum) currently has much the same depth distribution, growth form, and abundance as M. spicatum (Fig. 2, Table 1 ), we expanded the milfoil community to encompass coontail. Map segments in which the mean abundance of milfoil and coontail together exceeds the mean abundance of all other species combined and the water is less than 2.5 m deep are classified as shallow milfoil-coontail communities. The deepwater milfoil-coontail community is analogous but occurs in water that is 2.5 m or more deep. (Rooted macrophytes are not currently found deeper than 3.0 m in Wingra. ) All segments where the mean abundance of the two genera Nuphar and Nymphaea together exceeded that of all other plants are classified as a Nuphar-Nymphaea community. Remaining vegetated areas are classified as a mixed community. Results
Species abundances and community types The 1991 and 1992 surveys found V. americana, Chara spp., Potamogeton illinoensis Morong, Utricularia macrorhiza LeContc (previously called Utricularia vulgaris L. ), and the native milfoil M. sibiricum in Lake Wingra, all of which were absent in 1969 (Table 2). These species apparently returned to the lake some time in the early 1970s, as Chara was found in the 1970 survey (Table 2) and K americana, P. illinoensis, and U. macrorhiza were found among specimens collected in 1975 and 1976 (J. Titus, personal communication, 1992). Ranunculus longirostris Godron has not been found in the lake since the 1969 survey. The emergent species Scirpus validus Vahl. and Typha spp. were found in 1969 but not sampled in 1991-1992 because the minimum sampling depth of 0.5 m was outside their depth range. Over the 1969-1992 period, there was a marked decline in the relative and absolute frequency ofM. spicatum, which was somewhat offset by an increase
A.S. Trebitzet al. / AquaticBotany 46 (1993)325-340
331
Table 2 Change in the relative frequency of macrophytes in Lake Wingra from 1969 to 1992. Species are ordered by decreasing frequency in the 1969 survey. Species found with a relative frequency of less than 1% are categorized as 'other'. A dash indicates a species was not found, while blanks indicate species on which information is lacking. 'Other' species found were: Scirpus validus Vahl in 1969, Ranunculus longirostris Godron. in 1969, Typha angustifolia L. and Typha latifolia L. in 1969, Zannichellia palustris L. in 1970 and 1992, and Utricularia macrorhiza LeConte in 1991 and 1992 Species
Per cent relative frequency
Myriophfllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Potamogeton natans L. Potamogeton nodosus Poir. Ceratophyllum demersum L. Elodea canadensis Michx. Heteranthera dubia (Jacq.) MacM. Potamogeton crispus L. Potamogeton zosteriformis Fernald. Najasflexilis Rostk. & Schmidt Potamogeton foliosus Raf. Potamogeton richardsonii Rydb. Chara spp. Myriophyllum sibiricum Kom. Potamogeton illinoensis Morong. Vallisneria americana Michx.
1969
1970
1977
1991
1992
68.4 8.1 7.4 6.2 3.0 2.9 0.7 0.7 0.5 0.4 0.2 0.1 0.1
64.6 6.9 8.0 5.8 2.3 10.3 0.3 0.3 0.2 0.3 0.1 0.1 0.6
52.1 13.0 7.4 0.7 2.7 19.8
25.9 6.2 5.1 1.0 26.6 2.3 7.8 3.1 2.4 5.3 2.6 0.7 1.6
29.5 7.8 7.0 0.8 23.6 1.9 4.8 1.3 2.9 7.5 0.8 1.9 5.4
5.1
2.9
-
1.4
-
-
2.6
-
-
1.1
0.1
1.4 0.4
0.3 0.6
Other
1.2
Shannon-Wiener diversity index
1.28
1.26
2.25
2.17
Table 3 Change in the absolute frequency of abundant macrophytes in Lake Wingra over time Species
Myriophyllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Ceratophyllum demersum L.
Per cent absolute frequency 1969
1970
1977
1991
1992
93 11 10 4
78 8 10 12
91 22 13 32
57 14 11 58
51 13 12 41
in C. demersum (Tables 2 and 3, Fig. 3). M. spicatum and C. demersum now occur with roughly equal frequency. Total pondweeds (Potamogeton spp.) and waterlilies (Nymphaea and Nuphar spp. ) remained relatively constant over this period. Heteranthera dubia (Jacq.) MacM. and N. flexilis, present
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100
8O
[]
riD
other
[ ] N. flex#is [ ] H. dubia
~- 60 .m
[ ] C. demersum [ ] M. spicatum [ ] Potamogeton spp.
(D
.=>40 _m
•
20
I •
• Nuphar & Nymphaea
.I
0 1969 1970
1977
1991 1992
Fig. 3. Change in relative frequency of important species and plant groups over time.
[ ] no plants l Nuphar & Nymphaea [ ] mixed shallow milfoil-coontail 1[ ] deep milfoil-coontaU
~
~
Fig. 4. Plant community distribution in Lake Wingra in 1991 and 1992. Map segments were defined by depth contours and the location of survey transects.
but rare in 1969 and 1970, have expanded considerably (Fig. 3). The Shannon-Wiener diversity index (Pielou, 1977 ) has increased from about 1.3 in 1969 and 1970 to about 2.2 in 1991 and 1992 (Table 2). The shift away from milfoil dominance was also reflected in a changing
A.S. Trebitz et aL /Aquatic Botany 46 (1993) 325-340
333
balance of community types in the lake (Fig. 4 ). The shallow milfoil-coontail community in 1991 and 1992 occupied only 33% and 35% of segments respectively, versus 65% in 1969 (Nichols and Moil, 1971 ). The mixed community increased from 15% in 1969 to 21% in 1991 and 25% in 1992. The Nuphar-Nymphaea community varied only slightly (7% to 10% and 6%). The deep milfoil-coontail community appears to be quite variable ( 13% in 1969, 29% in 1991, 15% in 1992 ), largely because deeper littoral areas of the lake can change rapidly from sparse to no vegetation. Many segments that had a deep milfoil-coontail community in 1991 were bare in 1992 (Fig. 4). Interestingly, Nymphaea beds no longer appear to be as monospecific as reported by Nichols and Mori ( 1971 ). They rarely found Nymphaea mixed with any other species, but we invariably found a dense understory of submerged macrophytes in Nymphaea stands.
Biomass, cover, and distribution changes Total 1991 plant biomasses at the 1.5 m and 2.5 m depths across the three westernmost segments (419 _+85 g m - 2 and 534 + 214 g m - 2 respectively; mean _+SE) were not statistically distinguishable (two-tailed t-test) from the biomass of virtually pure M. spicatum stands in these same segments in 1970 ( 306 + 92 g m -2 and 140 __.79 g m -2). While the biomass data are quite variable, they do suggest that the current more diverse plant community has a biomass as great as, or greater than, the former milfoil-dominated one. Therefore, the loss of milfoil biomass during the late 1970s (Carpenter, 1980) has been compensated for by increases in other plants (largely coontail). Along with biomass, the coverage of M. spicatum was reduced in the late 1970s. Gustafson and Adams ( 1973 ) report only a slightly reduced coverage in the east end of the lake in 1971, and the 1977 distribution of milfoil was similar to that of the early 1970s (personal communication in Carpenter, 1980). However, milfoil in the southeast end of the lake thinned by 1978 (Carpenter, 1980), and had almost disappeared by 1979 (Jones et al., 1983 ). The current milfoil distribution largely matches that recorded in 1972 by Prentki (1979), although milfoil is still sparse on the southeast side of the lake (Fig. 5 ). The maximum depth to which milfoil grows appears to change quite rapidly. In 1991, milfoil was found to a depth of 3.0 m. In 1992, milfoil grew to a depth of only 2.0 or 2.5 m (Figs. 2 and 5 ) and only coontail grew as deep as 3.0 m. At mid-depths, milfoil rake indices in 1992 were generally higher than in 1991 (Fig. 5). We also examined the possibility of cover changes in the vegetation as a whole. Two of the 3.0 m depth segments were barren in 1991 and six were barren in 1992 (Fig. 4). Biomass quadrats were all vegetated at the 1.5 m depth in 1991, but half were bare at 2.5 m. Rake casts in 1991 and 1992 were similarly all vegetated at 1.5 m but 46% and 53% respectively, were bare at 2.5 m. Open areas at the shallower depths appear to be comparable in 1970,
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A.S. Trebitz et al. / Aquatic Botany 46 (1993) 325-340
dry weight (g/m 2) [ ] zero [ ] < 100 [ ] 100-200
[ ] 200-300 •
> 300
rake index [ ] zero ~<0.5
[ ] 0.5-1.0 [ ] 1.0-1.5 1>1.5
Fig. 5. The distribution of M. spicatum in 1970, 1991 and 1992. Direct comparisons between abundances derived from rake samples ( 1991-1992) and biomass quadrats (1970) are not valid, but these maps can reveal whether areas that were dense in 1970 remain so in 1991 and 1992. Data for deeper areas in 1970 are missing. 1991, and 1992. In 1970, 81% of the lake in the 0-2.4 m depth range was vegetated (Gustafson and Adams, 1973 ). O f the rake samples taken from a depth o f 0.5 to 2.5 m, 81% had vegetation in 1991 and 88% in 1992. In general, the outer edge of plant growth is variable, but vegetation coverage in areas less than 2.5 m deep appears to be relatively constant. To evaluate possible causes of the milfoil decline, we looked for correlations of the 1991 and 1992 M. spicatum distribution with depth and with 1970 sediment texture, organic content, or milfoil biomass. No significant relationships were found, except that in all 3 years, milfoil was most abundant at intermediate depths (Fig. 5 ). When segments were standardized for depth, there was again no significant correlation between 1991 and 1992 milfoil distribution and 1970 milfoil biomass or sediment characteristics. We were also unable to detect any spatial patterns (Figs. 5 and 6 ), such as m a r k e d changes in just one area o f the lake, which correlations might fail to reveal.
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A.S Trebitz et aL /Aquatic Botany 46 (1993) 325-340
% fine particles [7-~< 10
[ ] 10-15 [ ] 15-20 •
>20
% organic content []
DS-lO [] lo-15 Fig. 6. The distribution of sediment texture (percentage of fine particles) and organic content in 1970. Sediment data for deeper areas are unavailable.
Discussion
Patterns of species change While total plant cover in Lake Wingra remains similar to that found in the late 1960s and early 1970s, species diversity has increased dramatically and M. spicatum no longer dominates the community. This change is evident in the species frequencies and diversities (Tables 2 and 3, Fig. 3 ) and the distribution of community types (Fig. 4). It is tempting to attribute gains by native species to their release from competition by M. spicatum. Some species, notably P. illinoensis and K americana, are known to compete poorly with M. spicatum (Titus and Adams, 1979 ) and may have expanded as 3/1. spicatum declined. However, both these species were last reported in the lake in the 1920s (Nichols and Mori, 1971 ), well before M. spicatum made its appearance. The reports cited in the Lake history section suggest that the vegetation of Lake Wingra was depauperate prior to the arrival of milfoil. Furthermore, competition among aquatic macrophytes has not often been demonstrated in natural settings (Nichols, 1990; McCreary, 1991; Nichols, 1992 ). It is appropriate, then, to look further for factors which may have led to changes in the Lake Wingra flora. Jones et al. (1983) suggest that the late 1970s decline in milfoil biomass and distribution in Lake Wingra was associated with a decline in light penetration over the same period (Fig. 7 ). Indeed, C. demersum, Elodea canaden-
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336
0.0
E t-
0.4
0.8 ,v,
t--
o0
1.2 1972
1977
1982
1987
1992
Fig. 7. Recent trends in Lake Wingra Secchi disk transparency. Mean and range of the available summer data are plotted; open circles represent years for which only one sample was available. Data pertaining to 1970s are from Jones et al. (1983), 1986-1990 data are courtesy of the Wisconsin Self Help Lake Monitoring Program, and the 1991-1992 data are our own. The water surface is at the top of the left-hand axis.
sis Michx., Potamogeton crispus L., Potamogeton foliosus Raf., V. americana, and H. dubia, all have a positive association with turbid conditions (Nichols, 1992) and have expanded in Wingra, while Potamogeton nodosus Poir. has a negative association with turbidity and has declined. Photosynthetic efficiency curves for M. spicatum indicate that it is less shade tolerant than either V. americana, E. canadensis, or C. demersum (Guilizzoni, 1977; Titus and Adams, 1979; Madsen et al., 1991 ). Other factors, however, argue against the role of increasing turbidity as a factor in species change. Chara spp., P. illinoensis, Potamogeton zosteriformis Godron. and M. sibiricum are also negatively associated with turbidity (Nichols, 1992), yet have expanded in Lake Wingra. Also, laboratory growth studies are deceptive with respect to the light tolerance of M. spicatum, as it can compensate for low light conditions by concentrating its biomass at the water surface, thereby shading out competitors (Titus and Adams, 1979; Madsen et al., 1991 ). If light was in fact limiting in the late 1970s, it is no longer, as recent Secchi data (Fig. 7) show light penetration in the lake is now comparable to early 1970s levels. Whatever the reason, the late 1970s was clearly a time of stress for milfoil in Lake Wingra (Carpenter, 1980). Jones et al. (1983 ) suggested that phytoplankton and epiphytes compete for light with rooted macrophytes in Lake Wingra. Fitzgerald (1969) reported that Lake Wingra experienced heavy phytoplankton growths from the 1920s to 1950s, but was relatively clear in 1968 when milfoil was abundant. He also suggested that milfoil harvesting in the summer of 1968 led to phytoplankton increases and decreasing water clar-
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ity. Epiphyte growth on M. spicatum was quite heavy during the early 1970s (Gustafson and Adams, 1973; Carpenter, 1980), but was relatively minor in 1991 and 1992. We do not know whether patterns in algae growth are cause or consequence of changes in rooted vegetation. Additional evidence of stress is that milfoil apparently did not flower in 1977 (Carpenter, 1980), while we observed frequent flowering in 1991-1992. Other factors of the littoral environment may also be important to community change. Chara and P. illinoensis are both intolerant of disturbance (Davis and Brinson, 1980 ), and have only recently expanded in the lake (Table 2 ). Nicholson ( 1981 ) classifies C. demersum, E. canadensis, and H. dubia as highly resistant to physical disturbance such as harvesting, and I1. americana and P. crispus as moderately resistant. These species have increased in Lake Wingra, while Potamogeton spp., which are generally susceptible to harvesting (Nicholson, 1981 ), declined in the early 1970s (Table 2). However, while other Madison lakes were harvested extensively to control aquatic weeds, Lake Wingra was harvested routinely only in the vicinity of the boat launch and swimming area. Chara is susceptible to browsing by carp (King and Hunt, 1967), but their levels have been low since the late 1950s (Baumann et al., 1974). In general, no major changes to the littoral environment appear to have occurred.
Pattern and cause of milfoil decline Researchers have proposed that a variety of factors control M. spicatum invasions in lakes and reservoirs (reviewed by Carpenter, 1980; Painter and McCabe, 1988 ). We have already addressed light, turbidity and competition from other macrophytes and algae. We have no new data to permit us to directly evaluate other possible control agents. However, by examining changes in distribution patterns, we can at least speculate about which causes of milfoil decline are supported. Ifmilfoil responds to factors of the littoral environment which are independent of its presence (e.g. climate), it should retreat to formerly preferred areas as conditions become less favorable for its growth. Alternatively, ifM. spicatum locally depletes sediment resources, exhibits selfinhibition, or is subject to the density-dependent attention of a herbivore, parasite, or pathogen, milfoil should move about the littoral zone, not remaining abundant in the same place for very long. There is evidence that M. spicatum distribution is related to sediment characteristics. Laboratory work by Barko and Smart (1986) showed reduced milfoil growth on sediments of increasing organic matter, and Nichols ( 1971 ) found milfoil biomass in Lake Wingra in 1970 to be weakly correlated with sediment texture (clay and silt content). Unsuitable sediment texture and organics should develop independent of the presence of milfoil and so cause the M. spicatum distribution to contract to the most favorable areas. If, in-
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stead, possibly limiting sediment nutrients (such as nitrogen and phosphorus: Nichols and Keeney, 1976; Adams and Prentki, 1982; Smith and Barko, 1990) are depleted by milfoil, a shifting distribution pattern would be expected. Our data do not provide evidence for a change in milfoil distribution according to either of the proposed patterns. We were unable to find a relationship between the current milfoil distribution and sediment characteristics in 1970, either using correlations or by comparing high and low points on the maps. For example, sediments are particularly fine-textured in the east end of the lake (Fig. 6) but milfoil showed no greater tendency to change in these areas (Fig. 5 ). A retreat to or away from areas preferred in 1970 (Fig. 5 ) was not evident either. Our data may have been of insufficient resolution to detect such changes. Painter and McCabe (1988), investigating milfoil decline in Ontario lakes, were also unable to show a sediment effect on milfoil when grown in the laboratory on sediments taken from lakes which either still or no longer supported milfoil growth. We also looked for a pattern of milfoil decline near storm sewers entering the lake. Toxicants borne by runoff could suppress milfoil in the vicinity of storm sewer outlets (Carpenter, 1980). Also, light could have been limited by filamentous algae such as Oedogonium spp., which grew especially densely on milfoil near storm sewers in the early 1970s (McCracken et al., 1974). However, Carpenter (1980) noted no reduced milfoil growth near storm drains entering the west end of the lake, and neither did we. Our 1991-1992 surveys suggest that milfoil declined in a largely random pattern around the lake, failing to clearly support any of the discussed mechanisms for change. We are fairly certain that milfoil has not retreated to former preferred areas, but has remained growing (albeit more sparsely) over most of the littoral zone. Conclusions In Lake Wingra, as elsewhere, the natural invasion cycle of M. spicatum seems to be a sudden rise to dominance, followed by a sharp decline. However, the recovery of lakes following a milfoil decline has been poorly documented. Our work shows that rather than disappearing from Lake Wingra, M. spicatum has persisted at lower levels and become integrated into a more diverse plant community. Lakewide plant biomass and coverage declined when milfoil first 'crashed' but have now rebounded to former levels. Thus, native vegetation has expanded into areas formerly occupied by milfoil. To the best of our knowledge, these community changes have occurred in the absence of major changes in the trophic status, management, or use of the lake.
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339
Acknowledgments We thank Elizabeth Deppe and staff of the Wisconsin Department of Natural Resources and the Center for Limnology for sampling assistance. Thomas Martin and two anonymous reviewers provided helpful comments on the manuscript. Funding for this research was provided in part by the Federal Aid in Sport Fishing Restoration Act under Project F-95-P and the Wisconsin Department of Natural Resources. References Adams, M.S. and Prentki, R.T., 1982. Biology, metabolism, and function of littoral submersed weedbeds of Lake Wingra, Wisconsin, USA: a summary and review. Arch. Hydrobiol. Suppl., 62: 333-409. Barko, J.W. and Smart, R.M., 1986. Sediment-related mechanisms of growth limitation in submerged macrophytes. Ecology, 67:1328-1340. Baumann, P.C., Kitchell, J.F., Magnuson, J.J. and Kayes, T.F., 1974. Lake Wingra 1837-1973: A case history of human impact. Proc. Wise. Acad. Sci. Arts Lett., 62: 57-94. Carpenter, S.R., 1980. The decline of Myriophyllum spicatum in a eutrophic Wisconsin lake. Can. J. Bot., 58: 527-535. Davis, G.J. and Brinson, M.M., 1980. Response of submersed vascular plant communities to environmental change. Tcch. Rep. No. FWS-OBS-79-33, US Fish and Wildlife Service, KearneysviUe, WV, 69 pp. Deppe, E.R. and Lathrop, R.C., 1993. The aquatic macrophyte community of Lake Mendota in 1989-1990. Proc. Wise. Acad. Sci. Arts Lctt., in press. Drake, J.A., Mooney, H.A., di Castri, F., Groves, R.H., Kruger, F.J., Rejmanek, M. and Williamson, M. (Editors), 1989. Biological Invasions: a Global Perspective. John Wiley, New York, 525 pp. Elton, C.S., 1958. The Ecology of Invasions by Animals and Plants. Chapman and Hill, London, 181 pp. Fitzgerald, G.P., 1969. Some factors in the competition or antagonism among bacteria, algae, and aquatic weeds. J. Phycol., 5:351-359. Guilizzoni, P., 1977. Photosynthesis of the submergent macrophyte Ceratophyllum demersum in Lake Wingra. Proc. Wise. Acad. Sci. Arts Lctt., 65:152-162. Gustafson, T.D. and Adams, M.S., 1973. The remote sensing of aquatic macrophytes. Rep. No. 24, Institute of Environmental Studies, University of Wisconsin, Madison, 25 pp. Johnstone, I.M., 1986. Plant invasion windows: a time-based classification of invasion potential. Biol. Rev., 61: 369-394. Jones, C., Walti, K. and Adams, M.S., 1983. Phytoplankton as a factor in the decline of the submersed macrophyte Myriophyllum spicatum L. in Lake Wingra, Wisconsin, USA. Hydrobiologia, 107: 213-219. King, D.R. and Hunt, G.S., 1967. Effect of carp on vegetation in a Lake Erie marsh. J. Wildl. Manage., 31: 181-188. Lathrop, R.C., 1989. The abundance of aquatic macrophytes in the Yahara Lakes. Res. Manage. Find. No. 22, Wisconsin Department of Natural Resources, Madison, WI, 4 pp. Lind, C.T. and Cottam, G., 1969. The submerged aquatics of University Bay: a study in eutrophication. Am. Midl. Nat., 81: 353-369. Lodge, D.M., 1993. Species invasions and deletions: community effects and responses to eli-
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mate and habitat change. In: P.M. Karieva, J.G. Kingsolver and R.B. Huey (Editors), Biotic Interactions and Global Change. Sinauer Associates, Sunderland, MA, pp. 367-387. Madsen, J.D., Hartleb, C.F. and Boylen, C.W., 199 I. Photosynthetic characteristics of Myriophyllum spicatum and six submerged macrophyte species native to Lake George, N.Y. Freshwater Biol., 26: 233-240. McCracken, M.D., Gustafson, T.D. and Adams, M.S., 1974. Productivity of Oedogonium in Lake Wingra, Wisconsin. Am. Midl. Nat., 92: 247-254. McCreary, N., 1991. Competition as a mechanism of submersed macrophyte structure. Aquat. Bot., 41: 177-193. Nichols, D.S. and Keeney, D.R., 1976. Nitrogen nutrition ofMyriophyllum spicatum: variation of plant tissue nitrogen concentration with season and site in Lake Wingra. Freshwater Biol., 6: 137-144. Nichols, S.A., 1971. The distribution and control of macrophyte biomass in Lake Wingra. Ph.D. Thesis, University of Wisconsin, Madison, 111 pp. Nichols, S.A., 1990. Interspecific associations of some Wisconsin lake plants. Proc. Wisc. Acad. Sci. Arts Lett., 78:111-128. Nichols, S.A., 1992. Depth, substrate and turbidity relationships of some Wisconsin lake plants. Proc. Wisc. Acad. Sci. Arts Lett., 80:97-118. Nichols, S.A. and Mori, S., 1971. The littoral macrophyte vegetation of Lake Wingra: an example of a Myriophyllum spicatum invasion in a southern Wisconsin Lake. Proc. Wisc. Acad. Sci. Arts Lett., 59:107-119. Nicholson, S.A., 1981. Changes in submersed macrophytes in Chautauqua Lake, 1937-1975. Freshwater Biol., 11: 523-530. Painter, D.S. and McCabe, K.J., 1988. Investigation into the disappearance of eurasian watermilfoil from the Kawartha lakes. J. Aquat. Plant Manage., 26: 3-12. Pielou, E.C., 1977. Mathematical Ecology. John Wiley, New York, 385 pp. Prentki, R.T., 1979. Depletion of phosphorus from sediment colonized by Myriophyllum spicatum L. In: J.E. Breck, R.T. Prentki and O.L. Loucks (Editors), Proceedings of a Conference on Aquatic Plants, Lake Management, and Ecosystem Consequences of Lake Harvesting, Madison, WI, 14-16 February 1979, Institute of Environmental Studies, University of Wisconsin, Madison, pp. 161-176. Smith, C.S. and Barko, J.W., 1990. Ecology of Eurasian watermilfoil. J. Aquat. Plant Manage., 28: 55-64. Titus, J.E. and Adams, M.S., 1979. Coexistence and the comparative light relations of the submersed macrophytes Myriophyllum spicatum L. and Vallisneria americana Michx. Oecologia, 40: 273-286. Zimmermann, F.R., 1953. Waterfowl habitat survey and food habit studies, 1940-1943. Wisconsin Conservation Department Final Report, Pittmann-Robertson Project No. 6-R, 176 pp.
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0304-3770/93/$06.00 © 1993 - Elsevier Science Publishers B.V. All rights reserved
Patterns of vegetation change in Lake Wingra following a Myriophyllum spicatum decline Anett S. Trebitz *,a, Stanley A. Nichols b, Stephen R. Carpenter~, Richard C. Lathrop c "Centerfor Limnology, Universityof Wisconsin, 680 N. Park Sreet, Madison, W15 3 706, USA bWisconsin Geologicaland Natural History Survey, 3817 Mineral Point Road, Madison, W153705, USA CBureauof Research, WisconsinDepartment of Natural Resources, 3911 Fish Hatchery Road, Madison, W153711, USA (Accepted 15 June 1993)
AbsUa~
The invading aquatic plant, Myriophyllum spicatum L. is a management concern in many North American lakes because it replaces native species and because its dense growth can be a nuisance to lake users. It is common for M. spicatum to expand quickly upon reaching a lake, remain the most abundant littoral plant for a number of years, and then decline rather rapidly. This pattern held true for Lake Wingra, Dane County, WI, where M. spicatum dominated the littoral vegetation during the late 1960s, but abruptly declined during the 1970s. In this paper, we explore the changes in the Lake Wingra plant community that have occurred in the wake of the M. spicatum decline. We present results of 1991 and 1992 vegetation surveys indicating that M. spicatum, while no longer the dominant macrophyte, remains an important member of the Lake Wingra plant community. It and Ceratophyllum demersum L. now make up roughly equal parts of the littoral vegetation, and native species, rare or absent during the 1960s, are growing well. By comparing current plant distributions with those found earlier, we examined probable causes for the M. spicatum decline; no single factor seemed to be responsible. The miifoil decline in Lake Wingra has been sustained over roughly two decades while the native vegetation has expanded.
Introduction Species invasions and their effects are of considerable interest to ecologists (Elton, 1958; Johnstone, 1986; Drake et al., 1989; Lodge, 1993; and references therein). Often, the expansion of a disruptive invader is closely monitored, but once the invasion 'problem' diminishes, much less attention is given to the system's recovery. This pattern has-applied to the invasion of Eurasian milfoil (Myriophyllum spicatum L. ) in many North American lakes. It has been observed repeatedly that M. spicatum exhibits 'boom and bust' *Corresponding author.
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growth patterns, but it is not known what precipitates a decline. Proposed agents include nutrient depletion, light levels, competition, herbivory, parasites or pathogens, and management effects (Carpenter, 1980; Painter and McCabe, 1988; Smith and Barko, 1990). Often, competing factors cannot be separated because of insufficient information on lake conditions prior to the arrival of milfoil. There is also generally little information on plant community changes following a milfoil decline (Smith and Barko, 1990 ). In Lake Wingra, WI, we are fortunate to have records of macrophyte composition during the height of milfoil growth and for the first years of its decline ( 1969-1977 ). In this paper, we compare the results of 1991-1992 vegetation surveys with earlier data to describe how the vegetation has changed, approximately 15 years after milfoil began to decline. Probable causes of the decline are examined by comparing the present distribution of M. spicatum with patterns that might be expected from various causative agents. Lake history Lake Wingra is the smallest of several lakes within the Madison metropolitan area of Dane County, WI. This alkaline, eutrophic lake has a maximum depth of 4.3 m, a mean depth of 2.7 m, and a surface area of just under 140 ha (Baumann et al., 1974). About a third of its area is littoral (Adams and Prentki, 1982). Its shoreline is almost entirely within parks, while much of its watershed is residential. A few houses line its northern shore, as do a public swimming beach and a boat launch. Historically, Lake Wingra had a diverse plant community, with many floating and emergent species as well as submerged macrophytes. Prior to the 1900s, the area was marshy and the lake shore was extensively bordered by emergent plants (Baumann et al., 1974). Wild celery ( Vallisneria americana Michx. ) was reportedly quite abundant. Dredging and filling in the early 1900s and lowered water levels eliminated much of the marsh on the northeast and east sides of the lake. The carp (Cyprinus carpio L. ), abundant by 1915, also had a negative influence on vegetation (Baumann et al., 1974). Macrophyte abundances and diversities generally declined over the first half of the century, although Najas flexilis Rostk. & Schmidt, Potamogeton pectinatus L., Potamogeton natans L. and Ceratophyllum demersum L. were still common in 1943 (Zimmermann, 1953 ). V. americana was lost from the lake during this period. In the 1950s, carp control measures were taken and the macrophyte community recovered somewhat. At this time, the vegetation was dominated by pondweeds (Potamogeton spp. ), the native milfoil Myriophyllum sibiricum Kom. (formerly called Myriophyllum exalbescens Fernald), and coontail ( C. demersum) (Zimmermann, 1953 ). It is not known exactly when M. spicatum invaded Lake Wingra, but it must have expanded rapidly, since it was not reported prior to the 1960s, yet was
A.S. Trebitz et al. / Aquatic Botany 46 (1993) 325-340
327
abundant by 1966 (Nichols and Mori, 1971 ). A survey of the lake flora in 1969 documented the loss of M. sibiricum, V. americana, and several Potamogeton species along with the dominance ofM. spicatum (Nichols and Mori, 1971 ). Eurasian milfoil was also a problem in other Madison lakes during this time (Lind and Cottam, 1969 ) and resulted in a dramatic increase in the biomass of weeds (principally milfoil) cut and removed from the lakes (Lathrop, 1989 ). By the mid 1970s, milfoil had declined dramatically in all the Madison lakes for reasons that remain unclear (Carpenter, 1980; Lathrop, 1989). Methods
1991 and 1992 vegetation surveys We sampled Lake Wingra vegetation using a rapid survey method, whereby a weighted, double-headed rake is thrown into the water from a boat, dragged about 2 m across the bottom, and then pulled up by means of an attached line (Deppe and Lathrop, 1993 ). Plants recovered from the rake tines are identified to species and assigned an abundance ranging from 0 (not present) to 5 (very abundant). Abundances are based roughly on the biomass of plants of each species present in the sample. Sampling was done along transects perpendicular to the shore, with four replicate samples taken at each 0.5 m depth across the littoral zone. While this is not a method that yields precise abundance data, as is the case for example with sampling quadrats or transects by SCUBA, the rake method has the advantage of being much faster, and of covering more area, thereby increasing the chances of encountering rare species. Our 1991-1992 rake transects covered historically studied areas (Nichols and Mori, 1971 ) and avoided the swimming beach and boat launch along the northern shoreline (Fig. 1 ). The results of the two 1991 surveys show that the rake survey method is robust and repeatable: lakewide relative frequencies for the common species are quite similar (Table 1 ), and depth distributions of the dominant species in 1991 are comparable (Fig. 2). The 1991 surveys are in less agreement on the abundances of rare species, in part because our 16 August survey crew had less experience in identifying the rarer species. Also, rare species are usually more variable among surveys. We used our 6 August 1991 data for a comparison with frequencies from earlier years, but pooled the two 1991 surveys for distribution analyses of the common species. Limited biomass sampling was done in the west end of the lake concurrent with the 6 August 1991 rake survey. The total dry weight of all plants removed from 0.1 m 2 quadrats was recorded (plants dried for 48 h at 60°C). A total of 20 quadrats were sampled, ten taken randomly along each of the 1.5 and 2.5 m depth contours in the west end of the lake.
328
A.S. Trebitzet aL / AquaticBotany46 (1993)325-340
approximatelocationofsurveytransects ...................................... 6 A u g 9 1
IN ~
Fig. 1. Map of Lake Wingra showing location of survey transects. Depth contours (m) are approximate. Table 1 Relative frequency of macrophytes found by two 1991 rake surveys. The 6 August 1991 survey also found Potamogeton natans L., Potamogeton zosteriformis Fernald, Chara spp., Myriophyllum sibiricum Kom., Potamogeton illinoensis Morong, Vallisneria americana Michx. and Utricularia macrorhiza LeConte Species
Myriophyllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Ceratophyllum demersum L. Elodea canadensis Michx. Heteranthera dubia (Jacq.) MacM. Potamogeton crispus L. Najasflexilis Rostk. & Schmidt Potamogeton foliosus Raf. Potamogeton richardsonii Rydb.
Per cent relative frequency 6 Aug.
16 Aug.
25.9 6.2 5.1 26.6 2.3 7.8 3.1 5.3 2.6 0.7
24.0 5.8 7.9 27.1 2.7 9.1 2.1 10.0 2.1 5.2
Comparison with earlier data Lakewide relative frequency data were used to compare the plant community found in this study with that measured in 1969 (Nichols and Mori, 1971 ), 1970 (Nichols, 1971 ), and 1977 (Kimbel and Johnson, cited in Carpenter,
329
A.S. Trebitz et aL /Aquatic Botany 46 (1993) 325-340 6 Aug 91 © © 16 Aug 91 -. ...........• 18 Aug 92
,,,,"
',.
.~_
~2 V
8
H. dubia
g2 IV. flex#is .5
A
~
.
P. pectinatus
oT--
o.s
=
1.o
1.s 2.o 2.s water depth (m)
.
3.0
Fig. 2. The depth distribution of dominant plants (more than 5% relative abundance) in 1991 and 1992.
1980). Relative frequency data came from line-intercept surveys in 1969 and 1977, from quadrats in 1970, and from our rake surveys in 1991 and 1992. Since relative frequencies are computed from presence/absence information, comparison of these data is unaffected by differences among how the various methods measure abundance. The complete plant data set for 1977 has been lost, but species cited in Carpenter (1980) account for 97.1% of the relative frequency. Detailed measurements of lakewide milfoil biomass and sediment organic content and texture (percentage of particles less than 0.5 mm in diameter) were made in Lake Wingra in 1970 at the height of the milfoil invasion (Nichols, 1971 ). To compare 1970 milfoil biomasses with our own limited 1991 biomass samples, we computed 1970 biomass means and standard errors using only samples from approximately the 1.5 and 2.5 m depths in the west end of the lake. We also analyzed vegetation changes through correlation and
330
A.S. Trebitz et al. / Aquatic Botany 46 (1993) 325-340
map analyses that used spatial information in the 1970 and 1991-1992 data. We divided the lake into eight broad regions, chosen through a combination of the 1991-1992 transect coverage and the littoral zone topography (Fig. 1 ). Within each region, six 0.5 m depth classes were defined, resulting in a total of 48 map segments around the lake. Plant data for 1991-1992, and 1970 plant and sediment data were then assigned to these segments based on depth and position along the shoreline. Correlation analyses and distribution maps are based on the mean plant or sediment value within each segment. We also compared our 1991-1992 plant data with macrophyte community distributions established in 1969 by Nichols and Mori ( 1971 ). They identified five macrophyte communities in Lake Wingra: the shallow and deepwater milfoil communities, the Nuphar community, the Nymphaea community, and a mixed community. We adapted their classification to apply to conditions currently found in the lake. Because coontail (C. demersum) currently has much the same depth distribution, growth form, and abundance as M. spicatum (Fig. 2, Table 1 ), we expanded the milfoil community to encompass coontail. Map segments in which the mean abundance of milfoil and coontail together exceeds the mean abundance of all other species combined and the water is less than 2.5 m deep are classified as shallow milfoil-coontail communities. The deepwater milfoil-coontail community is analogous but occurs in water that is 2.5 m or more deep. (Rooted macrophytes are not currently found deeper than 3.0 m in Wingra. ) All segments where the mean abundance of the two genera Nuphar and Nymphaea together exceeded that of all other plants are classified as a Nuphar-Nymphaea community. Remaining vegetated areas are classified as a mixed community. Results
Species abundances and community types The 1991 and 1992 surveys found V. americana, Chara spp., Potamogeton illinoensis Morong, Utricularia macrorhiza LeContc (previously called Utricularia vulgaris L. ), and the native milfoil M. sibiricum in Lake Wingra, all of which were absent in 1969 (Table 2). These species apparently returned to the lake some time in the early 1970s, as Chara was found in the 1970 survey (Table 2) and K americana, P. illinoensis, and U. macrorhiza were found among specimens collected in 1975 and 1976 (J. Titus, personal communication, 1992). Ranunculus longirostris Godron has not been found in the lake since the 1969 survey. The emergent species Scirpus validus Vahl. and Typha spp. were found in 1969 but not sampled in 1991-1992 because the minimum sampling depth of 0.5 m was outside their depth range. Over the 1969-1992 period, there was a marked decline in the relative and absolute frequency ofM. spicatum, which was somewhat offset by an increase
A.S. Trebitzet al. / AquaticBotany 46 (1993)325-340
331
Table 2 Change in the relative frequency of macrophytes in Lake Wingra from 1969 to 1992. Species are ordered by decreasing frequency in the 1969 survey. Species found with a relative frequency of less than 1% are categorized as 'other'. A dash indicates a species was not found, while blanks indicate species on which information is lacking. 'Other' species found were: Scirpus validus Vahl in 1969, Ranunculus longirostris Godron. in 1969, Typha angustifolia L. and Typha latifolia L. in 1969, Zannichellia palustris L. in 1970 and 1992, and Utricularia macrorhiza LeConte in 1991 and 1992 Species
Per cent relative frequency
Myriophfllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Potamogeton natans L. Potamogeton nodosus Poir. Ceratophyllum demersum L. Elodea canadensis Michx. Heteranthera dubia (Jacq.) MacM. Potamogeton crispus L. Potamogeton zosteriformis Fernald. Najasflexilis Rostk. & Schmidt Potamogeton foliosus Raf. Potamogeton richardsonii Rydb. Chara spp. Myriophyllum sibiricum Kom. Potamogeton illinoensis Morong. Vallisneria americana Michx.
1969
1970
1977
1991
1992
68.4 8.1 7.4 6.2 3.0 2.9 0.7 0.7 0.5 0.4 0.2 0.1 0.1
64.6 6.9 8.0 5.8 2.3 10.3 0.3 0.3 0.2 0.3 0.1 0.1 0.6
52.1 13.0 7.4 0.7 2.7 19.8
25.9 6.2 5.1 1.0 26.6 2.3 7.8 3.1 2.4 5.3 2.6 0.7 1.6
29.5 7.8 7.0 0.8 23.6 1.9 4.8 1.3 2.9 7.5 0.8 1.9 5.4
5.1
2.9
-
1.4
-
-
2.6
-
-
1.1
0.1
1.4 0.4
0.3 0.6
Other
1.2
Shannon-Wiener diversity index
1.28
1.26
2.25
2.17
Table 3 Change in the absolute frequency of abundant macrophytes in Lake Wingra over time Species
Myriophyllum spicatum L. Potamogeton pectinatus L. Nuphar and Nymphaea spp. Ceratophyllum demersum L.
Per cent absolute frequency 1969
1970
1977
1991
1992
93 11 10 4
78 8 10 12
91 22 13 32
57 14 11 58
51 13 12 41
in C. demersum (Tables 2 and 3, Fig. 3). M. spicatum and C. demersum now occur with roughly equal frequency. Total pondweeds (Potamogeton spp.) and waterlilies (Nymphaea and Nuphar spp. ) remained relatively constant over this period. Heteranthera dubia (Jacq.) MacM. and N. flexilis, present
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100
8O
[]
riD
other
[ ] N. flex#is [ ] H. dubia
~- 60 .m
[ ] C. demersum [ ] M. spicatum [ ] Potamogeton spp.
(D
.=>40 _m
•
20
I •
• Nuphar & Nymphaea
.I
0 1969 1970
1977
1991 1992
Fig. 3. Change in relative frequency of important species and plant groups over time.
[ ] no plants l Nuphar & Nymphaea [ ] mixed shallow milfoil-coontail 1[ ] deep milfoil-coontaU
~
~
Fig. 4. Plant community distribution in Lake Wingra in 1991 and 1992. Map segments were defined by depth contours and the location of survey transects.
but rare in 1969 and 1970, have expanded considerably (Fig. 3). The Shannon-Wiener diversity index (Pielou, 1977 ) has increased from about 1.3 in 1969 and 1970 to about 2.2 in 1991 and 1992 (Table 2). The shift away from milfoil dominance was also reflected in a changing
A.S. Trebitz et aL /Aquatic Botany 46 (1993) 325-340
333
balance of community types in the lake (Fig. 4 ). The shallow milfoil-coontail community in 1991 and 1992 occupied only 33% and 35% of segments respectively, versus 65% in 1969 (Nichols and Moil, 1971 ). The mixed community increased from 15% in 1969 to 21% in 1991 and 25% in 1992. The Nuphar-Nymphaea community varied only slightly (7% to 10% and 6%). The deep milfoil-coontail community appears to be quite variable ( 13% in 1969, 29% in 1991, 15% in 1992 ), largely because deeper littoral areas of the lake can change rapidly from sparse to no vegetation. Many segments that had a deep milfoil-coontail community in 1991 were bare in 1992 (Fig. 4). Interestingly, Nymphaea beds no longer appear to be as monospecific as reported by Nichols and Mori ( 1971 ). They rarely found Nymphaea mixed with any other species, but we invariably found a dense understory of submerged macrophytes in Nymphaea stands.
Biomass, cover, and distribution changes Total 1991 plant biomasses at the 1.5 m and 2.5 m depths across the three westernmost segments (419 _+85 g m - 2 and 534 + 214 g m - 2 respectively; mean _+SE) were not statistically distinguishable (two-tailed t-test) from the biomass of virtually pure M. spicatum stands in these same segments in 1970 ( 306 + 92 g m -2 and 140 __.79 g m -2). While the biomass data are quite variable, they do suggest that the current more diverse plant community has a biomass as great as, or greater than, the former milfoil-dominated one. Therefore, the loss of milfoil biomass during the late 1970s (Carpenter, 1980) has been compensated for by increases in other plants (largely coontail). Along with biomass, the coverage of M. spicatum was reduced in the late 1970s. Gustafson and Adams ( 1973 ) report only a slightly reduced coverage in the east end of the lake in 1971, and the 1977 distribution of milfoil was similar to that of the early 1970s (personal communication in Carpenter, 1980). However, milfoil in the southeast end of the lake thinned by 1978 (Carpenter, 1980), and had almost disappeared by 1979 (Jones et al., 1983 ). The current milfoil distribution largely matches that recorded in 1972 by Prentki (1979), although milfoil is still sparse on the southeast side of the lake (Fig. 5 ). The maximum depth to which milfoil grows appears to change quite rapidly. In 1991, milfoil was found to a depth of 3.0 m. In 1992, milfoil grew to a depth of only 2.0 or 2.5 m (Figs. 2 and 5 ) and only coontail grew as deep as 3.0 m. At mid-depths, milfoil rake indices in 1992 were generally higher than in 1991 (Fig. 5). We also examined the possibility of cover changes in the vegetation as a whole. Two of the 3.0 m depth segments were barren in 1991 and six were barren in 1992 (Fig. 4). Biomass quadrats were all vegetated at the 1.5 m depth in 1991, but half were bare at 2.5 m. Rake casts in 1991 and 1992 were similarly all vegetated at 1.5 m but 46% and 53% respectively, were bare at 2.5 m. Open areas at the shallower depths appear to be comparable in 1970,
334
A.S. Trebitz et al. / Aquatic Botany 46 (1993) 325-340
dry weight (g/m 2) [ ] zero [ ] < 100 [ ] 100-200
[ ] 200-300 •
> 300
rake index [ ] zero ~<0.5
[ ] 0.5-1.0 [ ] 1.0-1.5 1>1.5
Fig. 5. The distribution of M. spicatum in 1970, 1991 and 1992. Direct comparisons between abundances derived from rake samples ( 1991-1992) and biomass quadrats (1970) are not valid, but these maps can reveal whether areas that were dense in 1970 remain so in 1991 and 1992. Data for deeper areas in 1970 are missing. 1991, and 1992. In 1970, 81% of the lake in the 0-2.4 m depth range was vegetated (Gustafson and Adams, 1973 ). O f the rake samples taken from a depth o f 0.5 to 2.5 m, 81% had vegetation in 1991 and 88% in 1992. In general, the outer edge of plant growth is variable, but vegetation coverage in areas less than 2.5 m deep appears to be relatively constant. To evaluate possible causes of the milfoil decline, we looked for correlations of the 1991 and 1992 M. spicatum distribution with depth and with 1970 sediment texture, organic content, or milfoil biomass. No significant relationships were found, except that in all 3 years, milfoil was most abundant at intermediate depths (Fig. 5 ). When segments were standardized for depth, there was again no significant correlation between 1991 and 1992 milfoil distribution and 1970 milfoil biomass or sediment characteristics. We were also unable to detect any spatial patterns (Figs. 5 and 6 ), such as m a r k e d changes in just one area o f the lake, which correlations might fail to reveal.
335
A.S Trebitz et aL /Aquatic Botany 46 (1993) 325-340
% fine particles [7-~< 10
[ ] 10-15 [ ] 15-20 •
>20
% organic content []
DS-lO [] lo-15 Fig. 6. The distribution of sediment texture (percentage of fine particles) and organic content in 1970. Sediment data for deeper areas are unavailable.
Discussion
Patterns of species change While total plant cover in Lake Wingra remains similar to that found in the late 1960s and early 1970s, species diversity has increased dramatically and M. spicatum no longer dominates the community. This change is evident in the species frequencies and diversities (Tables 2 and 3, Fig. 3 ) and the distribution of community types (Fig. 4). It is tempting to attribute gains by native species to their release from competition by M. spicatum. Some species, notably P. illinoensis and K americana, are known to compete poorly with M. spicatum (Titus and Adams, 1979 ) and may have expanded as 3/1. spicatum declined. However, both these species were last reported in the lake in the 1920s (Nichols and Mori, 1971 ), well before M. spicatum made its appearance. The reports cited in the Lake history section suggest that the vegetation of Lake Wingra was depauperate prior to the arrival of milfoil. Furthermore, competition among aquatic macrophytes has not often been demonstrated in natural settings (Nichols, 1990; McCreary, 1991; Nichols, 1992 ). It is appropriate, then, to look further for factors which may have led to changes in the Lake Wingra flora. Jones et al. (1983) suggest that the late 1970s decline in milfoil biomass and distribution in Lake Wingra was associated with a decline in light penetration over the same period (Fig. 7 ). Indeed, C. demersum, Elodea canaden-
A.s. Trebitz et aL / Aquatic Botany 46 (1993) 325-340
336
0.0
E t-
0.4
0.8 ,v,
t--
o0
1.2 1972
1977
1982
1987
1992
Fig. 7. Recent trends in Lake Wingra Secchi disk transparency. Mean and range of the available summer data are plotted; open circles represent years for which only one sample was available. Data pertaining to 1970s are from Jones et al. (1983), 1986-1990 data are courtesy of the Wisconsin Self Help Lake Monitoring Program, and the 1991-1992 data are our own. The water surface is at the top of the left-hand axis.
sis Michx., Potamogeton crispus L., Potamogeton foliosus Raf., V. americana, and H. dubia, all have a positive association with turbid conditions (Nichols, 1992) and have expanded in Wingra, while Potamogeton nodosus Poir. has a negative association with turbidity and has declined. Photosynthetic efficiency curves for M. spicatum indicate that it is less shade tolerant than either V. americana, E. canadensis, or C. demersum (Guilizzoni, 1977; Titus and Adams, 1979; Madsen et al., 1991 ). Other factors, however, argue against the role of increasing turbidity as a factor in species change. Chara spp., P. illinoensis, Potamogeton zosteriformis Godron. and M. sibiricum are also negatively associated with turbidity (Nichols, 1992), yet have expanded in Lake Wingra. Also, laboratory growth studies are deceptive with respect to the light tolerance of M. spicatum, as it can compensate for low light conditions by concentrating its biomass at the water surface, thereby shading out competitors (Titus and Adams, 1979; Madsen et al., 1991 ). If light was in fact limiting in the late 1970s, it is no longer, as recent Secchi data (Fig. 7) show light penetration in the lake is now comparable to early 1970s levels. Whatever the reason, the late 1970s was clearly a time of stress for milfoil in Lake Wingra (Carpenter, 1980). Jones et al. (1983 ) suggested that phytoplankton and epiphytes compete for light with rooted macrophytes in Lake Wingra. Fitzgerald (1969) reported that Lake Wingra experienced heavy phytoplankton growths from the 1920s to 1950s, but was relatively clear in 1968 when milfoil was abundant. He also suggested that milfoil harvesting in the summer of 1968 led to phytoplankton increases and decreasing water clar-
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ity. Epiphyte growth on M. spicatum was quite heavy during the early 1970s (Gustafson and Adams, 1973; Carpenter, 1980), but was relatively minor in 1991 and 1992. We do not know whether patterns in algae growth are cause or consequence of changes in rooted vegetation. Additional evidence of stress is that milfoil apparently did not flower in 1977 (Carpenter, 1980), while we observed frequent flowering in 1991-1992. Other factors of the littoral environment may also be important to community change. Chara and P. illinoensis are both intolerant of disturbance (Davis and Brinson, 1980 ), and have only recently expanded in the lake (Table 2 ). Nicholson ( 1981 ) classifies C. demersum, E. canadensis, and H. dubia as highly resistant to physical disturbance such as harvesting, and I1. americana and P. crispus as moderately resistant. These species have increased in Lake Wingra, while Potamogeton spp., which are generally susceptible to harvesting (Nicholson, 1981 ), declined in the early 1970s (Table 2). However, while other Madison lakes were harvested extensively to control aquatic weeds, Lake Wingra was harvested routinely only in the vicinity of the boat launch and swimming area. Chara is susceptible to browsing by carp (King and Hunt, 1967), but their levels have been low since the late 1950s (Baumann et al., 1974). In general, no major changes to the littoral environment appear to have occurred.
Pattern and cause of milfoil decline Researchers have proposed that a variety of factors control M. spicatum invasions in lakes and reservoirs (reviewed by Carpenter, 1980; Painter and McCabe, 1988 ). We have already addressed light, turbidity and competition from other macrophytes and algae. We have no new data to permit us to directly evaluate other possible control agents. However, by examining changes in distribution patterns, we can at least speculate about which causes of milfoil decline are supported. Ifmilfoil responds to factors of the littoral environment which are independent of its presence (e.g. climate), it should retreat to formerly preferred areas as conditions become less favorable for its growth. Alternatively, ifM. spicatum locally depletes sediment resources, exhibits selfinhibition, or is subject to the density-dependent attention of a herbivore, parasite, or pathogen, milfoil should move about the littoral zone, not remaining abundant in the same place for very long. There is evidence that M. spicatum distribution is related to sediment characteristics. Laboratory work by Barko and Smart (1986) showed reduced milfoil growth on sediments of increasing organic matter, and Nichols ( 1971 ) found milfoil biomass in Lake Wingra in 1970 to be weakly correlated with sediment texture (clay and silt content). Unsuitable sediment texture and organics should develop independent of the presence of milfoil and so cause the M. spicatum distribution to contract to the most favorable areas. If, in-
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stead, possibly limiting sediment nutrients (such as nitrogen and phosphorus: Nichols and Keeney, 1976; Adams and Prentki, 1982; Smith and Barko, 1990) are depleted by milfoil, a shifting distribution pattern would be expected. Our data do not provide evidence for a change in milfoil distribution according to either of the proposed patterns. We were unable to find a relationship between the current milfoil distribution and sediment characteristics in 1970, either using correlations or by comparing high and low points on the maps. For example, sediments are particularly fine-textured in the east end of the lake (Fig. 6) but milfoil showed no greater tendency to change in these areas (Fig. 5 ). A retreat to or away from areas preferred in 1970 (Fig. 5 ) was not evident either. Our data may have been of insufficient resolution to detect such changes. Painter and McCabe (1988), investigating milfoil decline in Ontario lakes, were also unable to show a sediment effect on milfoil when grown in the laboratory on sediments taken from lakes which either still or no longer supported milfoil growth. We also looked for a pattern of milfoil decline near storm sewers entering the lake. Toxicants borne by runoff could suppress milfoil in the vicinity of storm sewer outlets (Carpenter, 1980). Also, light could have been limited by filamentous algae such as Oedogonium spp., which grew especially densely on milfoil near storm sewers in the early 1970s (McCracken et al., 1974). However, Carpenter (1980) noted no reduced milfoil growth near storm drains entering the west end of the lake, and neither did we. Our 1991-1992 surveys suggest that milfoil declined in a largely random pattern around the lake, failing to clearly support any of the discussed mechanisms for change. We are fairly certain that milfoil has not retreated to former preferred areas, but has remained growing (albeit more sparsely) over most of the littoral zone. Conclusions In Lake Wingra, as elsewhere, the natural invasion cycle of M. spicatum seems to be a sudden rise to dominance, followed by a sharp decline. However, the recovery of lakes following a milfoil decline has been poorly documented. Our work shows that rather than disappearing from Lake Wingra, M. spicatum has persisted at lower levels and become integrated into a more diverse plant community. Lakewide plant biomass and coverage declined when milfoil first 'crashed' but have now rebounded to former levels. Thus, native vegetation has expanded into areas formerly occupied by milfoil. To the best of our knowledge, these community changes have occurred in the absence of major changes in the trophic status, management, or use of the lake.
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Acknowledgments We thank Elizabeth Deppe and staff of the Wisconsin Department of Natural Resources and the Center for Limnology for sampling assistance. Thomas Martin and two anonymous reviewers provided helpful comments on the manuscript. Funding for this research was provided in part by the Federal Aid in Sport Fishing Restoration Act under Project F-95-P and the Wisconsin Department of Natural Resources. References Adams, M.S. and Prentki, R.T., 1982. Biology, metabolism, and function of littoral submersed weedbeds of Lake Wingra, Wisconsin, USA: a summary and review. Arch. Hydrobiol. Suppl., 62: 333-409. Barko, J.W. and Smart, R.M., 1986. Sediment-related mechanisms of growth limitation in submerged macrophytes. Ecology, 67:1328-1340. Baumann, P.C., Kitchell, J.F., Magnuson, J.J. and Kayes, T.F., 1974. Lake Wingra 1837-1973: A case history of human impact. Proc. Wise. Acad. Sci. Arts Lett., 62: 57-94. Carpenter, S.R., 1980. The decline of Myriophyllum spicatum in a eutrophic Wisconsin lake. Can. J. Bot., 58: 527-535. Davis, G.J. and Brinson, M.M., 1980. Response of submersed vascular plant communities to environmental change. Tcch. Rep. No. FWS-OBS-79-33, US Fish and Wildlife Service, KearneysviUe, WV, 69 pp. Deppe, E.R. and Lathrop, R.C., 1993. The aquatic macrophyte community of Lake Mendota in 1989-1990. Proc. Wise. Acad. Sci. Arts Lctt., in press. Drake, J.A., Mooney, H.A., di Castri, F., Groves, R.H., Kruger, F.J., Rejmanek, M. and Williamson, M. (Editors), 1989. Biological Invasions: a Global Perspective. John Wiley, New York, 525 pp. Elton, C.S., 1958. The Ecology of Invasions by Animals and Plants. Chapman and Hill, London, 181 pp. Fitzgerald, G.P., 1969. Some factors in the competition or antagonism among bacteria, algae, and aquatic weeds. J. Phycol., 5:351-359. Guilizzoni, P., 1977. Photosynthesis of the submergent macrophyte Ceratophyllum demersum in Lake Wingra. Proc. Wise. Acad. Sci. Arts Lctt., 65:152-162. Gustafson, T.D. and Adams, M.S., 1973. The remote sensing of aquatic macrophytes. Rep. No. 24, Institute of Environmental Studies, University of Wisconsin, Madison, 25 pp. Johnstone, I.M., 1986. Plant invasion windows: a time-based classification of invasion potential. Biol. Rev., 61: 369-394. Jones, C., Walti, K. and Adams, M.S., 1983. Phytoplankton as a factor in the decline of the submersed macrophyte Myriophyllum spicatum L. in Lake Wingra, Wisconsin, USA. Hydrobiologia, 107: 213-219. King, D.R. and Hunt, G.S., 1967. Effect of carp on vegetation in a Lake Erie marsh. J. Wildl. Manage., 31: 181-188. Lathrop, R.C., 1989. The abundance of aquatic macrophytes in the Yahara Lakes. Res. Manage. Find. No. 22, Wisconsin Department of Natural Resources, Madison, WI, 4 pp. Lind, C.T. and Cottam, G., 1969. The submerged aquatics of University Bay: a study in eutrophication. Am. Midl. Nat., 81: 353-369. Lodge, D.M., 1993. Species invasions and deletions: community effects and responses to eli-
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