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Key words: warm-water lake, seasonal plankton succession. Abstract. Common seasonal ...... VanDolah, F. M., D. L. Roelke & R. M. Greene, 2001. Health.

Hydrobiologia 513: 205–218, 2004. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.


Interannual variability in the seasonal plankton succession of a shallow, warm-water lake Daniel Roelke, Yesim Buyukates, Mike Williams & Jason Jean Texas A&M University, Wildlife and Fisheries Sciences, 2258 TAMUS, College Station, TX 77843-2258, U.S.A. Tel: 979-845-0169. Fax: 979-845-4096. E-mail: [email protected] Received 11 February 2003; in revised form 23 July 2003; accepted 6 August 2003

Key words: warm-water lake, seasonal plankton succession

Abstract Common seasonal plankton succession patterns in temperate lakes are well understood, and were described in the popular PEG-model. Seasonal plankton succession in warm-water lakes, however, is not as well known. Recent theory suggests that some lake systems are characteristic of having alternate system-states, where one of the system-states is characterized by dominance of cyanobacteria, and transition between system-states can be abrupt and undeterminable. Lake Somerville, a shallow, well-mixed, warm-water reservoir located in eastern TX, U.S.A., experiences occasional periods of cyanobacteria dominance. To increase our understanding of seasonal plankton dynamics in warm-water systems, we analyzed 14-years of plankton data spanning a 22-year period. During this period, succession dynamics characteristic of those described by the PEG-model were observed, as well as succession dynamics expected during periods of cyanobacteria dominance, i.e., greater accumulated phytoplankton biovolume, low secondary productivity, and low light penetration. In addition to the PEG-model and cyanobacteria type system-states, other states of the system that were intermediate between these were observed. Therefore, we conclude the lake does not behave according to the alternate system-states model. The change from year to year in early-year cyanobacteria dominance was abrupt and non-monotonic during this period. In addition, the early year performance of cyanobacteria appeared to influence the plankton succession trajectory for the remainder of the season. While the magnitude of lake-flushing early in the year accounted for ∼37% of variability in cyanobacteria prevalence, many of the traditional factors impacting cyanobacteria dominance appeared insignificant.

Introduction Efforts to improve water quality and protect human health often focus on phytoplankton biology and factors that influence the dynamics of phytoplankton succession (Paerl, 1988b; Reynolds, 1994; Roelke & Buyukates, 2001, 2002; VanDolah, et al., 2001). In lakes these efforts are assisted through use of various predictive models that are founded on our understanding of foodweb linkages and interactions between biota and the abiotic environment. One such model, coined the PEG-model, details the seasonal succession of phytoplankton in temperate lakes. In this model, early spring is dominated by highly edible, rapidly growing, r-selected species.

As the season progresses this community gives way to less edible, slower growing, k-selected species. Factors that influence the succession pattern include limitation by multiple nutrients and preferential grazing (Sommer et al., 1986; Reynolds, 1988; Sommer, 1989b; Sterner, 1989). The seasonal succession pattern culminates with the onset of fall holomixis followed by ice cover (Horne & Goldman, 1994; Kalff, 2002). Our knowledge of seasonal succession patterns and causative mechanisms is not as extensive for warmwater lakes as it is for temperate lakes. While some trends observed in warm-water lakes are consistent with descriptions in the PEG-model, deviations are known to occur. For example, in some systems winter

206 is not a time of low phytoplankton biomass, and nutrients can become limiting before onset of stratification (Grover et al., 1999). In addition, seasonal succession often terminates through a combination of increased lake-flushing and decreased temperature during the fall and winter months (Paerl, 1988a; Pollingher et al., 1998; Scheffer, 1998). Another model, complementary to the PEG-model, describes various competing system-states that dictate the type of seasonal succession that will occur in a given year (Scheffer et al., 1997; Scheffer, 1998; Carpenter et al., 1999, 2001). For example, a system-state of high water-quality might exist that is characteristic of low water column nutrients, low phytoplankton, high secondary productivity, and higher water transparency. Succession dynamics in this system-state would likely be similar to dynamics described by the PEG-model. In the same system, however, another system-state of low water-quality might also occur that is characteristic of high water column nutrients, dominance of cyanobacteria, low secondary productivity, and low light penetration. Plankton dynamics in this system-state would be very unlike dynamics described in the PEG-model. Both system-states appear to be self-sustaining over the course of a season. For example, through non-selective and highly effective grazing some cladocerans help to maintain conditions of deeper light penetration, which tends to select against many forms of cyanobacteria because of the greater light penetration to mixing depth ratio (Scheffer, 1998). In turn, this may promote phytoplankton communities of higher quality to consumers. The resulting increased zooplankton productivity then helps to keep phytoplankton biomass cropped, and maintains deeper light penetration. Conversely, dominance of cyanobacteria forms often results in diminished grazing rates and grazer populations (Schindler, 1971; Gliwicz & Lampert, 1990; Haney et al., 1994), allowing accumulation of phytoplankton biomass and causing decreased light penetration, a condition which favors continued growth of cyanobacteria over other phytoplankton (Scheffer, 1998). Factors that influence which system-state occurs in a lake are debated, and include sediment phosphorus content, magnitude of flushing, and the ratio of light penetration to mixing depth (Scheffer et al., 1997; Scheffer, 1998; Carpenter et al., 1999, 2001). In theory, transitional values for these factors must exist where a lake system is situated on the edge of competing system-states. In other words, which system-state

is selected in a given year becomes very sensitive to environmental conditions early in the year. Indeed, this has been documented in some lakes, where the lake abruptly changed back and forth between the system-states (Scheffer, 1998). Anecdotal evidence from Lake Somerville, TX, suggested that the lake might behave in accordance with the alternate system-states model. For example, an abrupt change from the typical plankton succession pattern to dominance of cyanobacteria occurred in 1998, although no system level-disturbances that might have caused this change were observed (Khan, pers. commun.). The lake reverted back to typical succession in 1999, again with no accompanying systemlevel changes observed. System behavior of this nature is consistent with a lake situated on the edge of competing system-states. Our purpose here is to test this hypothesis, as well as to increase our understanding of plankton succession patterns in warm-water lakes. We use both newly collected data and historical data that combined represent 14-years of plankton data spanning a period of 22-years. We structure our analyses within the framework of the PEG and alternate system-states models.

Materials and methods Lake Somerville is a reservoir located in eastern Texas constructed by the Army Corps of Engineers in 1967 for purposes of flood control. Typically, the lake’s surface area is ∼11 460 acres, mean depth is ∼4 m, and maximum depth is ∼9 m. For this research, water quality and plankton data were analyzed from one station located at the lower end of the lake, 30◦ 19.08 latitude, 96◦ 31.31 longitude. Our recent sampling period occurred over 4-years (1999–2002) with trips conducted monthly. The time of sampling ranged between 9:30 a.m. and 10:30 a.m. for all trips. We also compiled 11-years of historical data from this station (1980–1990, winter, spring, and summer sampling), as described below. Sampling of the plankton was from surface waters only, i.e., ∼0.5 m depth. Phytoplankton and zooplankton were collected using a sampling bottle and 12-l Schindler trap, respectively. Phytoplankton samples were preserved in 2% glutaraldehyde, and zooplankton samples were preserved in 2% buffered formalin. Both were enumerated using inverted lightmicroscopy (Utermohl, 1958). For phytoplankton, appropriate dimensions were measured and biovolume

207 calculated using formulas of corresponding geometric shapes (Wetzel & Likens, 1991), and identification was at least to the taxonomic level of genus (Prescott, 1978). Zooplankton was categorized into copepods (adults and nauplii), cladocerans (Daphnia and Bosmina), rotifers, and ciliates. Historical phytoplankton data were made available through the Army Corps of Engineers with identification at least to the taxonomic level of genus. Water-quality parameters were also determined at the time of plankton sampling, which included nutrients, temperature and light penetration. Nutrient samples were filtered through 47 mm GF/F filters and frozen for transport to the laboratory. Nutrient analyses included nitrate (NO3 ), nitrite (NO2 ), ammonia (NH4 ), urea, soluble reactive phosphorus (SRP), and silicate (SiO2 ) using autoanalyzer methodology (Armstrong & Sterns, 1967; Harwood & Kuhn, 1970). Temperature was measured using a water quality multiprobe (Hydrolab H20), and profiles were approximated by collecting data at 0.5-m (towards surface) and 1-m intervals. Light penetration was estimated by taking the secchi depth. Samples for determination of chlorophyll a and phaeophytin a were also collected. Typically, 50 ml samples were taken in triplicate and filtered onto 47 mm GF/F filters using gentle vacuum, then frozen for transport to the laboratory. Chlorophyll a and phaeophytin a samples were extracted in 10 ml 90% acetone and analyzed using a fluorometric method (APHA, 1998). Lake flushing was determined for a 20-year period (1983–2002) spanning the historical and more recent periods of study. Daily records of total inflow and reservoir volume were obtained from the Army Corps of Engineers and U.S. Geological Survey. Flushing was calculated by dividing the total inflow by the lake volume. The mean flushing rate over a period of 90days prior to each of the sampling dates was also calculated. In addition to investigating phytoplankton community structure, we analyzed diversity at the taxonomic level of genus. Diversity was calculated using the Shannon–Weaver index: x  pi log 2 (pi ), (1) H = i=1

where pi was the proportion of biovolume of genera i relative to the total biovolume on a given date, and x was the total number of phytoplankton genera observed.

Linear and nonlinear relationships were investigated between flushing and light penetration early in the year with cyanobacteria dominance using linear, power, and logarithmic regression curve fitting (Synergy & Software, 1999). We used the mean 90-day flushing values when attempting to correlate flushing with cyanobacteria dominance. For the analysis focusing on light penetration, winter secchi depths were plotted against spring cyanobacteria dominance.

Results Seasonal hydrology in Lake Somerville for the period of 1983 through 2002 was erratic (Fig. 1a). In some years flushing was greatest in the fall and winter months, but in other years maximum flushing occurred in the summer or spring. Periods of high flushing were few and short-lived, typically lasting only 2–3-days. Flushing magnitude reached as high as ∼0.35 d−1 , and on average was 0.004 ± 0.011 d−1 . Focusing on the period of 1999 through 2002 (Fig. 1b), it can be seen that late 1999 through mid-year 2000 was a relatively dry period. The mean flushing for the winter period of January through March was 0.0015 ± 0.0023 d−1 . Conversely, late 2000 through 2001 was a relatively wet period, with flushing for the same winter period averaging 0.0082 ± 0.0106 d−1 . Winter flushing during late 2001 through mid-2002 was intermediate, and had a mean value of 0.0021 ± 0.0028 d−1 . There is a chance that we underestimated lakeflushing because we used the entire volume of the lake in our calculation. For example, if flow through the lake were confined to areas near the historical riverbed flushing in this portion of the lake would be much higher. We do not believe this was the case, however. The historical riverbed runs from west to east, and the predominant winds are south to north. Indeed, windrows spanning the width of the lake are frequently observed orthogonal to the riverbed, which are indicative of Langmuir circulation churning the lake. Temperature trends in Lake Somerville for the period spanning late 1999 through 2002 showed a range of ∼20 ◦ C, with minimum and maximum values of ∼10 and ∼30 ◦ C, respectively (Fig. 2). Because Lake Somerville is shallow, there was no consequential seasonal thermal stratification, and the lake was subject to frequent and complete mixing events. In our analyses below we use secchi depth as a surrogate for


Figure 1. Flushing in Lake Somerville, TX for a 20-year period (a), which spanned the more recent (b) and historical periods of study. Consistent seasonal trends were not apparent. The period from late-1999 through 2000 was relatively dry, 2001 was relatively wet, and 2002 was intermediate.

secchi depth to mixing depth ratio, because mixing depth in Lake Somerville was constant, i.e., the entire water column. In the more recent data, some seasonal succession trends consistent with the PEG-model were observed in Lake Somerville, but only during 2001 and 2002. In these years a spring phytoplankton bloom occurred dominated by diatoms (Fig. 3a). Chain-forming Melosira spp. dominated the diatom community during the spring bloom of 2001 resulting in decreased

diversity. During the spring bloom of 2002 multiple diatom species prospered and diversity was maintained (Figs 3b and 4a). The spring 2001 bloom of phytoplankton appeared to be highly edible, despite the large size of the Melosira chains, which were typically between 300 and 600 µm in length (see Sterner, 1989). Evidence supporting this notion were the dramatic increases in rotifer, cladoceran, and copepod populations (Fig. 5a, b) that were synchronous with a decrease in phyto-


Figure 2. Temperature profiles in Lake Somerville, TX for the late-1999 through 2002 period. Thermal stratification was inconsequential during this period and complete mixing of the water-column was frequent.

plankton biovolume (Fig. 3a). In addition, a spike in the phaeophytin a to chlorophyll a ratio, which is an indicator of grazing activity (Mantoura & Llewellyn, 1983; Roelke et al., 1997), shortly followed the zooplankton population peaks (Fig. 5c). Finally, gut contents of gizzard shad, Dorosoma cepedianum, were filled with Melosira spp. at this time (Fejes et al., 2003). While no grazing experiments were conducted during this research, we feel that the phytoplankton and zooplankton population demographics, and coincident changes in the phaeophytin a to chlorophyll a ratio are strong indicators of active grazing. The spring 2002 bloom of phytoplankton appeared to be less edible compared to 2001. Cyanobacteria abundance, which included Microcyctis spp., a genera known to inhibit grazers (see Horne & Goldman, 1994), was greater in 2002 (Figs 3a and 4b). In addition, rotifer, cladoceran, and copepod populations were not as abundant, and the phaeophytin a to chlorophyll a ratio was not as high (Fig. 5). The impact of grazing activity, which we deduced from observed population demographics and phaeophytin a to chlorophyll a ratio, in the spring of 2001 and 2002 altered the water-column environment. The deepest secchi depths observed in this study occurred after the termination of these spring blooms, which we now refer to as clear-water phases (Fig. 3c). Inorganic nutrients increased during the clear-water phases, presumably due to grazer egestion and excretion processes (Sommer et al., 1986; Sommer, 1989b;

Sterner, 1989), with the increase quite pronounced in 2001 (Fig. 6). The increase in DIN was most dramatic during 2001, with almost all of it in the form of NO3 (Fig. 6a). The DIN to SRP ratio spiked at this time as well (Fig. 6c). In 2002, again DIN increased more than SRP, with most of the nitrogen in the form of NH4 . It appears that the demise of the spring bloom of diatoms in 2001 was not just a function of heavy grazing activity. SiO2 concentrations dropped below 3 µM (Fig. 6d), which is below the k S value for Si-limited growth for many diatoms (Horne & Goldman, 1994). Consequently, it is likely that SiO2 became limiting in Lake Somerville in the spring of 2001. During the winter and early spring, DIN and SRP did not appear to be limiting, as ambient concentrations were much higher than typical kS values for N-and P-limited growth. Zooplankton populations declined during the clear-water phases, but did not totally disappear, i.e., rotifers persisted (Fig. 5a, b). This allowed for a second phytoplankton bloom to occur that was more diverse and was comprised of a mix of genera that represented edible and ‘less’ edible forms (Figs 3 and 4). For example, based on cell-size, morphology, nutritional content, and results from previous research, taxa belonging to the genera Clamydomonus, Chlorella, Scenedesmus, and Crucigenia, as well as forms of centric diatoms, are deemed edible and are even considered preferred prey items, while taxa belonging to Microcystis are thought of as ‘less’ edible (Sommer


Figure 3. Phytoplankton bulk community structure (a) and diversity (b), and secchi depth (c). Phytoplankton biovolume was much greater in 2000 compared to 2001 and 2002, and was dominated by cyanobacteria. In 2001 a spring bloom dominated by diatoms was followed by a clear-water phase that gave way to a second more diverse bloom. By late summer cyanobacteria dominated the phytoplankton community. In 2002 a spring bloom occurred, again rich with diatoms but also cyanobacteria. A clear-water phase occurred followed by increased prevalence of cyanobacteria late in the summer. For the years 2001 and 2002, arrows indicate the times of zooplankton population and phaeophytin a to chlorophyll a ratio maxima.


Figure 4. Phytoplankton community structure at the taxonomic level of genus divided into diatoms (a), cyanobacteria (b), and green algae (c). During 2000, cyanobacteria dominated the phytoplankton community, and showed population shifts from dominance of Microcystis spp. in the winter to co-dominance of Aphanocapsa spp. and Aphanizominon spp. during the spring, and finally co-dominance of Aphanocapsa spp. and Oscillatoria spp. in the summer (note that the y-axis is scaled differently for panel b). During 2001, Melosira spp. dominated the spring bloom. This bloom gave way to a clear-water phase that was followed by a second more diverse bloom. Towards the late summer Phormidium spp. became dominant. In 2002, the spring bloom was comprised of multiple diatom species, and Microcystis spp., which persisted through the clear-water phase. Again, towards late summer Phormidium spp. became dominant. For the years 2001 and 2002, arrows indicate the times of zooplankton population and phaeophytin a to chlorophyll a ratio maxima.


Figure 5. Zooplankton community structure divided into rotifers, ciliates, and copepod nauplii (a) and cladocerans and adult copepods (b), and the phaeophytin a to chlorophyll a ratio as an indicator of grazing activity (c). Population maxima occurred for all zooplankton following the spring bloom in 2001 and 2002, although smaller peaks occurred in 2002. Heavy grazing activity in 2001, as indicated by the elevated phaeophytin a to chlorophyll a ratio, contributed to the clear-water phase that occurred at that time. During 2000, populations of large zooplankton were non-existent.

et al., 1986; Sterner, 1989; Sterner & Hessen, 1994; Horne & Goldman, 1994). As the summer progressed the community eventually shifted to dominance of cyanobacteria from the genus Phormidium. This pattern was more pronounced in 2001 than in 2002 (Figs 3a and 4b). The succession pattern in 2000 was very different from 2001 and 2002, and inconsistent with the PEG-model. During 2000 cyanobacteria dominated the phytoplankton community most of the time (Fig. 3a). Early in the spring Microcystis spp. dominated, and this population appeared to be carry-over from the autumn and winter of 1999 (Fig. 4b). Opposite of what was observed in 2001 and 2002, phytoplankton biovolume continued to increase through the spring months and into the early summer, and reached concentrations ∼2-fold greater than what was observed in 2001 and 2002. Cyanobacteria continued to dominate, but the community shifted to dominance of species

from the genera Aphanocapsa and Aphanizominon. A centric diatom (Stephanodiscus-like) also became prevalent. Seasonal SiO2 reached its minimum at this time, but was not as depleted as it was in 2001 (Fig. 6). Similarly, rotifer populations increased during this time, similar to 2002 but not to the same degree as in 2001, and larger zooplankton were scarce (Fig. 5a, b). A dramatic decrease in the phytoplankton biovolume occurred in the early summer of 2000, and it is unclear what might have caused it. It is not likely that grazing activity was a significant factor because there were no corresponding increases in grazer populations, the phaeophytin a to chlorophyll a ratio was low (Fig. 5), and there were no increases in nutrient concentrations (Fig. 6). It is not likely that pathogens significantly contributed either because signs of infection during microscopic examination were not detected. The phytoplankton quickly recovered, however, and increased in biovolume through the remaining summer months and into the autumn. Again, biovolume was ∼2-fold higher in 2000 compared to 2001 and 2002 during this period. Cyanobacteria continued to dominate the phytoplankton community, and the composition again shifted with Aphanocapsa spp. and Oscillatoria spp. sharing dominance (Figs 3a and 4b). Through the autumn and winter months the phytoplankton biovolume diminished (Fig. 3a), rotifer populations increased slightly (Fig. 5a), and nutrient concentrations increased (Fig. 6). Light penetration, as estimated from secchi depth, varied between 2000, 2001, and 2002. All years showed a similar pattern of deeper light penetration in the spring and early summer months, and less in the late summer and autumn months (Fig. 3c). But secchi depths were much less during the late-winter and spring in 2000, e.g., light penetration in 2000 was only about half of that in 2001. The historical data shows no consistent trend with pertinent aspects of the PEG-model. For example, a diatom-dominated spring bloom was not apparent in 1982 and 1989. Instead, green algae (Tetradron spp.) and cyanobacteria (Oscillatoria spp.) co-dominated in 1982, and a euglenoid (Trachelomonas spp.) dominated 1989 (Table 1). In addition, diatoms of the genera Melosira, a meroplanktonic strategist that is expected to become dormant during summer months, often persisted into the summer as a major contributor to the total phytoplankton biovolume. Also, summertime prevalence of cyanobacteria, observed in all years of our recent sampling, was not always observed in the historical data. For example, cyanobacteria only


Figure 6. Inorganic nutrient concentrations divided into species of nitrogen (a), soluble reactive phosphorus (b), N:P ratio (c), and silicate (d). Remineralization of all nutrients occurred in 2001 and 2002 during the clear-water phase, although less so in 2002. Nitrogen was returned to the water column at greater proportions than phosphorus because of the presence of cladocerans, which have high demand for phosphorus. The spring bloom of diatoms became silica-limited only in 2001, which contributed to the demise of the spring bloom in that year.


Figure 7. A nonlinear relationship between early-year lake flushing and early-year cyanobacteria prevalence was observed using a power function regression curve, which accounted for ∼37% of the variance. Note that many intermediate values were observed regarding the percent cyanobacteria composition.

accounted for 5%, 2%, and 19% of the total phytoplankton biovolume in 1982, 1988, and 1989, respectively. Finally, decreased light penetration as the season progresses from spring to summer, again observed in all years of our more recent sampling, was seldom observed in the historical data. The 3-years of monthly sampling in Lake Somerville indicated that cyanobacteria prevalence in the winter and spring does occur, another phenomenon not consistent with the PEG-model, but not every year. Historical data support this observation. For example, 51%, 76% and 41% of the phytoplankton biovolume was comprised of cyanobacteria in the winter samplings of 1980, 1988 and 1990, respectively (Table 1). Only once in the historical data was the springtime cyanobacteria contribution significant, i.e., 32% in 1982. A nonlinear relationship between the early season mean 90-d flushing and cyanobacteria prevalence was observed (Fig. 7), i.e., a power-function fit indicated that 37% of the variance in cyanobacteria prevalence could be explained by variations in lake

flushing. No significant relationships, linear or nonlinear, were detected between winter secchi depth and spring cyanobacteria prevalence. Discussion During 2001 the plankton succession pattern in Lake Somerville followed predictions made by the PEGmodel and observations from other warm-water lakes. For example, a spring bloom occurred which was dominated by diatoms. This is an early-stage of succession as described in the PEG-model for temperate lakes (Sommer et al., 1986; Sommer, 1989a), and is also a common occurrence in warm-water lakes (Porter et al., 1996; Phlips et al., 1997; Grover et al., 1999). In addition, it appeared that the stimulation of secondary productivity was a result of the dominance of highly edible phytoplankton forms, and that a coupling of SiO2 -limition and heavy grazing losses was the mechanism resulting in the rapid decrease of diatom biovolume. Both are well-documented events in lakes (Sommer et al., 1986; Sommer, 1989b; Horne & Gold-

215 Table 1. Historical records. Dates of sampling, seasonal categorization (w – winter, sp – spring, su – summer), secchi depth,% of total phytoplankton biovolume comprised of cyanobacteria, and phytoplankton genera comprising >10% of total phytoplankton biovolume with percent contribution in parentheses 1980



1983 1984







30-Jan 21-May 29-Aug 6-Mar 30-Apr 24-Aug 18-Feb 2-June 12-Aug 25-Jan 4-May 5-Jan 2-May 1-Aug 9-Jan 23-Apr 1-Aug 4-Mar 14-May 14-Aug 26-Jan 16-Apr 19-Aug 26-Jan 26-Apr 19-Jul 20-Jan 2-May 16-Aug 10-Jan 17-Apr 18-Jul

w sp su w sp su w sp su w sp w sp su w sp su w sp su w sp su w sp su w sp su w sp su

1.0 0.9 0.6 1.0 0.7 0.8 0.6 0.9 0.7 0.9 0.4 1.0 0.5 0.6 0.6 0.7 0.7 0.5 1.1 1.1 0.6 0.5 0.5 0.3 0.7 0.7 0.8 1.1 0.8 0.8 0.6 0.6

51% 3% 89% 17% 1% 65%

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