Pathways
Leading To Water Clarity Restoration In Ventura Marsh Following Benthivorous
Fish Removal
Exotic, benthivorous fish are among the major determinates of water quality in shallow, nutrient rich lakes. Of the exotic benthivores, perhaps the most widely distributed is the common carp (Cyprinus carpio). Common carp are native to Eurasia but have become established in North America, Europe, India, South Africa and Australia (Roberts et al. 1995). Populations of common carp are found throughout the continental United States with the lakes and rivers of the Midwest having the greatest densities (Courtney et al. 1984).
Common carp have been implicated in eutrophication of water bodies. The impact of common carp is largely attributed to its benthic feeding activities (Welcomme 1984). Potentially, carp may enhance phosphorus concentration (Breukelaar et al. 1994, Havens 1991, Brabrand et al. 1990, Lamarra 1975, Vanni and Findlay 1990), increase phytoplankton biomass, increase turbidity (Scheffer 1998) and reduce the abundance of submerged macrophytes (Crivelli 1983, Roberts et al. 1995, Barko and Smart 1981, Skubinna et al. 1995).
One important technique used to restore water quality in freshwaters is biomanipulation. Biomanipulation is a lake management tool aimed at increasing water clarity by manipulating the biomass of fish (Perrow 1997). Biomanipuations have been conducted throughout Europe and North America during the past fifty years, many of which have been successful in improving water clarity and/or lowering phytoplankton biomass (Drenner and Hambright 1999).
Despite the many studies there are still discrepancies as to the mechanism through which water clarity increases following fish biomanipulation. Benthivores structure aquatic systems through many processes and it is not completely certain which processes are involved in the increase in water clarity following biomanipulation.
Following benthivorous fish removal, water clarity may increase due to a reduction in suspended sediment. Benthivores feed on invertebrates, such as chironomids, oligochaeta, and mollusca, inhabiting the sediment by sucking in sediment and catching the invertebrates in their gill rackers. This feeding process forms small pits in the lake bottom and resuspends sediment (Scheffer 1998). In Lake Bleiswijkse and Lake Noorddiep benthivorous fish had a substantial effect on turbidity due to bioturbation of the sediment (Meijer et al. 1990). Experimental ponds stocked with common carp (Cyprinus carpio) showed a positive relationship between fish biomass and suspended solids (Breukelaar et al. 1994, Lougheed et al. 1998). A benthivorous fish density of 600 kg/ha in a shallow lake may reduce Secchi disk transparency to 0.4 m solely due to sediment resuspension (Meijer et al. 1990). Changes in water clarity following benthivorous fish removal may be due to reduced bioturbation of the sediment during foraging.
Reduction of phytoplankton biomass as a result of phosphorus limitation may be an alternate pathway through which water clarity increases following removal of benthivorous fish (Boers et al. 1991). Benthivores may increase phosphorus concentrations due to recycling of nutrients from the sediment and fish excretion (Breukelaar et al. 1994, Havens 1991, Brabrand et al. 1990, Lamarra 1975, Vanni and Findlay 1990). Phosphorus concentrations decreased in some biomanipulations involving a reduction of benthivores (Meijer et al. 1989, Hanson and Butler 1994, Meijer and Hosper 1997), although, other manipulations observed no change in nutrient concentrations (VanDonk et al. 1990, Lougheed 2000, Bonneau 1999).
Increased water clarity, due to a reduction in phytoplankton biomass, may be a result of trophic cascading. The idea being that a reduction in fish biomass will shift the zooplankton community from smaller-bodied to larger-bodied zooplankters, such as Daphnia. An increase in larger-bodied zooplankton may reduce phytoplankton biomass, leading to increased water clarity since larger-bodied zooplankton are more efficient grazers than smaller-bodied zooplankters (Brooks and Dodson 1965, Shapiro and Wright 1984, Carpenter et al. 1985). A shift from smaller-bodied to larger-bodied zooplankters may be expected for two reasons following removal of benthivores. First, removal of benthivores may reduce the predation pressure on larger-bodied zooplankton since juvenile benthivores selectively feed on the larger zooplankters (Drenner and Hambright 1999) and many adult benthivores switch between benthic and pelagic feeding (Lammens and Hoogenboezem 1991). Secondly, turbity may also structure the zooplankton community (Bonneau 1999, Lougheed and Chow-Fraser 1998). It has been shown that under conditions of high suspended sediment grazing by large cladocerans is hindered, leading to a zooplankton community consisting of smaller-bodied individuals (Kirk 1991, Hart 1988). If predation and/or suspended sediment decreases following removal of benthivores, the zooplankton community may shift towards larger-bodied zooplankton. Therefore, the reduction in fish biomass may have a cascading effect down through the aquatic trophic levels and eventually reduce the phytoplankton biomass leading to increased water clarity.
Expansion of macrophyte beds has been observed following biomanipulations involving removal of benthivores (Meijer et al. 1990, Ozimek et al. 1990, Hanson and Butler 1994, Meijer and Hosper 1997). This may occur since the higher water clarity allows for macrophytes to establish at greater depths (Barko and Smart 1981, Skubinna et al. 1995) and the uprooting of vegetation by foraging benthivores, such as carp, is reduced (Crivelli 1983).
Macrophytes growth during the clear water phase following fish removal can provide positive feedbacks, which help stabilize the clear water phase. Macrophytes help maintain high water clarity by competing with algae for nutrients and light (Perrow et al. 1997, Van Donk et al. 1993), providing refugia for zooplankton (Timms and Moss 1984, Schriver et al. 1995), increasing sedimentation of suspended particles (James and Barko 1990), and perhaps suppressing algal growth by allelopathy (Wium-Anderson et al. 1982).
The purpose of this study was to examine the impacts of fish removal, primarily benthivorous, from the temperate, eutrophic Ventura Marsh. Venture Marsh is unique in that it is a large, shallow, windy system located in a predominately agricultural watershed. This paper will describe the effectiveness of benthivorous fish removal in increasing water clarity of Ventura Marsh and attempt to elucidate the pathway by which water clarity is affected.
The western bay of Clear Lake (Little Lake), to which Ventura Marsh is
a tributary, was monitored as reference site for this study (Fig. 1). The Little Lake was chosen because it is
similar to Ventura Marsh in nutrient regime, size (127 ha), depth (mean depth
1.13 m), and is exposed to the same seasonal variability.
To determine the mechanism(s) by which benthivorous fish removal influenced Ventura Marsh, we studied water quality, plankton, benthos, and macrophytes before and after a series of fish kills. Initially, the fish removal was planned to occur at the end of the summer of 1999 sampling season, so that we could obtain a year of pre- and post-biomanipulation data. Difficulties in obtaining a substantial fish removal resulted in three attempts at rotenone application before a large removal was attained. The IDNR applied rotenone aerially to Ventura Marsh on August 17, 1999 and June 7, 2000. In addition, rotenone was applied under the ice on February 13, 2000. Water quality, plankton, and benthic communities were sampled every two weeks from April through October of 1999 and every two weeks from March through September of 2000 with a higher frequency of sampling employed following summer fish removals. Water quality and phytoplankton samples were collected from the Little Lake every two weeks on the same days as Ventura Marsh.
We assessed water quality variables at three sites on Ventura Marsh (Fig. 2) and at one site on the Little Lake. Dissolved oxygen, temperature, pH, and conductivity were measured at each half-meter interval in the water column. At each half-meter interval, we also collected water samples for analysis of total nitrogen, nitrate, ammonia, total phosphorus, silica and total, inorganic and volatile suspended solids. Total nitrogen and nitrate were analyzed using the second derivative method (Crumpton et al. 1992). The remaining analyses were conducted according to standard methods (American Public Health Association 1998). Secchi disk readings were also taken at these three sites.
Plankton and benthos samples were collected in excess. In August of 1999 we expanded our sampling of plankton and benthos from three samples to 13 planktons samples and 7 benthic samples (Fig. 2). Fewer benthic samples were taken because we felt that the distribution of benthos was more homogeneous than that of the plankton community. We counted a number of randomly chosen samples that was sufficient to yield an inter-replicate standard error of ≤ 20% of the mean (Downing 1979). After counting many of the 1999 plankton samples it was noticed that only 4 phytoplankton and 3 zooplankton samples were counted for most dates. We therefore reduced the number of samples collected in 2000 to 6 phytoplankton and 5 zooplankton samples. We continued to collect 7 benthic samples in 2000.
The phytoplankton samples were comprised of equal volumes of water taken from each half-meter interval, and were preserved with Lugol’s solution (APHA 1998). Samples were concentrated and sub-sampled with a Hensen-Stempel pipette. The volume of the sub-sample varied between 2 – 5 ml depending on density of cells. Using an inverted microscope, we identified, counted, and measured phytoplankton. Samples were counted until the most abundant species reached 125 except when Oscillatoria was most abundant. When Oscillatoria was most abundant, samples were counted until Oscillatoria reached 1000. Fifty cells of each taxa were measured from each sample except Oscillatoria, for which 250 cells were measured. Phytoplankton was identified to genus, with the exception of small cyanophyceae, using the keys of Ward and Whipple (1959) and Whitford and Schumacher (1984). Phytoplankton cells were measured and biomass estimated by applying basic geometric formulas (Findenegg 1974). From this information we calculated the biomass of edible (<30 mm) and inedible (≥30 mm) phytoplankton (Watson et al. 1992).
We used a 30 L Shindler-Patalus trap with a 61μm mesh net to collect zooplankton samples from the onset of the study until May 23, 2000, at which time we began sampling using a 61μm mesh Wisconsin net. We switched to a Wisconsin net because we were unable to submerge the Shindler-Patalus trap without disturbing the sediment during periods of low water in 2000. To determine the difference in efficiency between these two sampling devices, both collection methods were used simultaneously on three sampling dates. In terms of biomass, the Schindler-Patalus trap was found to be approximately 5% more efficient than the Wisconsin net. Therefore, zooplankton values from Wisconsin net samples were corrected such that they express the expected biomass, had the Schindler-Patalus trap been used. Zooplankton samples were preserved in 5% Formalin solution with 20 g/L of sucrose, and were later transferred to 70% ethanol. Samples were sub-sampled using a Hensen-Stempel pipette to obtain a volume with a minimum of 60 organisms (McCauley 1984). Using a stereomicroscope with 50X magnification, we identified, counted, and measured zooplankton. Twenty-five individuals of each taxa were measured. Rotifera and Cladocera were identified to genus and Copepoda to suborder using the key of Pennak (1989). We estimated zooplankton biomass by applying length-weight equations (Rosen 1981, Dumont et al. 1975) with the exception of Keratella spp. The biomass of Keratella spp. was estimated from biovolume (Ruttner-Kolisko 1977) assuming a specific gravity of 1.0 and a wet to dry weight ratio of 0.05 (Schindler and Noven 1971).
Grazing rate of cladocerans and
rotifers was approximated using abundance data and estimated filtering rates
from the literature (Grosselain et al. 1996, Haney 1973, Mourelatos and Lacroix
1990, Bogdan and Gilbert 1982).
Copepods were not included in the estimation of grazing rates. Only harpacticoid
and cyclopoid copepods were present in this study. Harpacticoids feed from the bottom (Pennack 1989) so they would
not contribute to the grazing of the phytoplankton in suspension. Cyclopoid grazing on phytoplankton is not
well understood (Adrian 1991) and therefore were not included in the estimated
grazing rate.
The sediment of Ventura Marsh is organic mud so we were able to obtain 1 L benthic samples with an Ekman grab. We filtered the sediment samples through a 600 µm sieve, and the portion remaining in the sieve was preserved in 5% Formalin solution with 20 g/L sucrose and 100 mg/L of rose bengal (Mason and Yevich 1967). We counted and identified all benthic organisms in the sample using the keys of Pennak (1989) and Merritt and Cummins (1996). The first 25 Chironomids and 30 oligochates in each were sample measured. Dry masses of Diptera and Gastropoda were estimated using length-weight equations (Benke et al. 1999, Eckblad 1970). We estimated oligochaete dry mass based on biovolume (Smit et al. 1993). The density of benthos was determined by dividing biomass by the volume of the sediment sample.
In order to evaluate the impact of fish manipulation on submerged macrophytes, we conducted macrophyte surveys in July 1999 and August 2000. Twelve north south transects evenly spread throughout the open water were surveyed in 1999. We recorded the species present along these transects. In 2000 the open water of the marsh was surveyed for submerged macrophytes with 27 north south transects approximately 20 m apart. A one-meter square quadrat was placed every 20 meters along these transects. We identified the species in the quadrats and estimated the percent cover of each species.
To gauge the success of the rotenone
treatments, the Iowa Department of Natural Resources conducted gill netting
surveys. Three gill nets, measuring 160
feet long with 2½ in bar mesh, were placed for 24 hours in Ventura Marsh on
four occasions (8/3/99, 9/10/99, 4/12/00, and 6/21/00). The fish captured were
identified and enumerated. During the last three gill net surveys carp were
categorized as either small (< 1.8 kg) or large
(≥ 1.8 kg). There was
reason to believe that the carp population was becoming re-established in late
summer of 2000, so electroshocking was conducted on September 1, 2000. Two transects were shocked for 7 minutes
each. The fish captured during electroshocking were identified and the lengths
of the first 50 fish of each species were measured.
Direct comparisons of water quality
values between the four time periods (pre-manipulation, postmanipulation 1,
postmanipulation 2 and postmanipulation 3) could be confounded by seasonal
variability. To reduce this, the
before-after-control-impact (BACI) method of analysis (Smith et al. 1993) was
used to determine whether differences in nutrient concentrations, suspended
solids, Secchi disk transparency and phytoplankton biomass were statistically
significant among the four time periods.
The Little Lake was used as the reference system in this analysis.
Prior to rotenone applications, gill
nets placed overnight in Ventura Marsh collected 170 common carp (Cyprinus carpio) (Table 1). After the first and second rotenone
applications 113 and 84 common carp were collected respectively. Of the carp captured following the first
rotenone application 54% were large fish (1.8 kg), while only 11% of the carp
captured following the second rotenone application were large. Only 2 common
carp, less than 1.8 kg in weight, were collected following the third
application in June of 2000, indicating that the majority of the carp
population had been eradicated by the three rotenone applications. Thirty-one bullheads (Ameiurus melas) were collected after the first rotenone application
but none were captured on the other collection dates. On September 1, 2000 two transects in the marsh were
electroshocked. A total of 73 fishes
(68 common carp, 4 buffalo, and 1 bullhead) were captured in transect one. Only
seven fishes (6 common carp and 1 buffalo) were captured in transect 2. The carp were primarily juveniles with an
average length of 15.7 cm (6.2 in) and a range of
11.9 – 26.7 cm (4.7 – 10.5 in).
Secchi disk transparency was generally quite low in the marsh (~0.35 m) but was significantly higher following the third rotenone application (BACI, p < 0.05). The highest Secchi disk transparency of 1.0 m was recorded on July 13, 2000 (Fig. 3), 6 weeks following the final rotenone treatment. The Secchi disk transparencies from May 10, 2000 to July 19, 2000 were slightly underestimated since the Secchi disk reached the marsh bottom at one or more of the sites without disappearing from view. The period following the third rotenone application is therefore referred to the clear water phase and the period prior as the turbid phase.
There was weak evidence that total phosphorus concentrations were reduced in the period following the third fish removal compared to the premanipulation period (BACI, p < 0.058). In the turbid phase the total phosphorus of Ventura Marsh was on average 147 μg/L higher than the total phosphorus concentration of the Little Lake, whereas in the clear water phase the difference was an average of 32 μg/L. In the turbid phase total phosphorus and phytoplankton biomass were strongly correlated (r2 = 0.73) but were less so in the clear water phase (r2 = 0.21) (Fig. 4).
Ammonia concentrations following the third rotenone application were significantly different from the premanipulation and postmanipulation 1 period (BACI, p < 0.05). Total nitrogen, nitrate, and silica showed no significant change in concentrations among the four treatment periods (BACI, p < 0.05).
Inorganic suspended solids (ISS) concentrations in the water column did not differ significantly among the four treatment periods (BACI, p < 0.05), although the lowest concentration of ISS occurred during the clear water phase (Fig. 3). Linear regression of Secchi disk depth and inorganic suspended solids showed a negative correlation (r2 = 0.51) (Fig. 5).
Volatile suspended solids following the third rotenone application were significantly lower than during the other three treatment periods (BACI, p < 0.05). Volatile suspended solids include both phytoplankton and organic detritus. No significant difference in phytoplankton biomass was detected between the four periods (BACI, p = 0.05). A sizeable decrease in phytoplankton biomass occurred between May 23, 2000 and June 27, 2000 (Fig. 6). Secchi disk transparency showed a strong negative linear relationship with phytoplankton biomass across the turbid and clear phases (r2 = 0.55, 0.74).
Phytoplankton
biomass steadily increased throughout 1999 and a similar trend began in 2000
(Fig. 6). However, after the third fish removal the phytoplankton biomass
decreased. Total phytoplankton biomass
and cyanophycea biomass appeared to be higher in 1999. Cyanophycea, Oscillatoria, and Actinastrum
were more prominent in 1999, while Closterium, Mersimopedia and Synedra
were more predominant in 2000.
Changes were observed in the zooplankton community across the fish manipulations (Fig. 7). The zooplankton biomass composition during the premanipulation period (4/12/99 – 8/16/99) was variable. There was a peak of Bosmina in early June and an increase in biomass of Chydorus during August that remained high until the first rotenone application. Cyclopoid copepods and Keratella biomass were fairly constant during this period. Following the first rotenone application, zooplankton biomass was low. There was a small peak of Brachionus followed by a steady biomass of Keratella, Chydorus, Cyclopoid copepods and nauplii. Chydorus was very abundant during this period. The biomass composition of zooplankton remained fairly similar throughout the postmanipulation 2 period (8/17/99 – 10/15/99). The zooplankton biomass in postmanipulation 2 was primarily cyclopoida copepods. In the spring, cyclopoid nauplii and Daphnia were present but remained very low throughout the rest of the postmanipulation 2 period. The zooplankton biomass composition underwent substantial changes following the third biomanipulation. The first zooplankter to increase following the third fish removal was Brachionus. Brachionus began to decrease in late June 2000, at which time the biomass of Daphnia, Ceriodaphnia, cyclopoid copepods, and nauplii began to increase. Approximately one month later the biomass of Bosmina increased. In early August Daphnia and Ceriodaphnia populations began to decline. By mid-August and September of 2000 the zooplankton community consisted mainly of Brachionus, Bosmina, cyclopoid copepods and nauplii.
Changes in the size distribution of cladocerans during this study reflect changes in the prominent cladoceran (Fig. 8). In 1999, the length of most cladocerans was between 0.2 – 0.3 mm, whereas, in 2000, a larger range of cladoceran lengths was observed. In 1999, Chydorus was the primary cladoceran while larger cladocerans, such as Daphnia and Ceriodaphnia, were more prominent in 2000. The distribution of copepod lengths remained similar throughout the study with a median length of approximately 0.6 mm (Fig. 8).
Estimated grazing rates of cladocerans and rotifers peaked in September 1999, June 2000, and July 2000 (Fig. 9). The 1999 peak corresponded to a high biomass of Chydorus. The June 2000 peak was due primarily to rotifers, while the July 2000 peak corresponds to high abundance of Daphnia and Ceriodaphnia. The July 2000 peak in grazing was the highest, with nearly 140% of the marsh water being filtered each day. There was little correlation between grazing rate and phytoplankton biomass when viewed across both the turbid and clear phase (r2 = 0.16, 0.23).
Oligochaetes and chironomids were the primary benthic organisms in Ventura Marsh (Fig. 10). The composition of the benthic community was similar throughout the study but the biomass changed considerably. In the premanipulation and postmanipulation 1 periods the average biomass of benthos was 22 and 32 mg/L respectively, whereas in postmanipulation 2 and 3 the average biomass was 95 and 116 mg/L. The distribution of oligochaete length remained similar throughout the study while the median length of Chironomids increased in the postmanipulation 3 period (Fig. 11).
Throughout this study cattails, Typha, surrounded the shoreline of Ventura Marsh and clumps of Lemna were present throughout the open water. In 1999, only six of the twelve transects were found to contain submerged macrophytes. Three species of submerged macrophytes were observed: sago pondweed (Potamogeton pectinatus), coontail (Ceratophyllum) and water lily (Nymphaea sp.). Sago pondweed and coontail were the primary submerged macrophytes with only one occurance of water lily. All submerged macrophytes in 1999 were found within 5 m of the shoreline. In the 2000 survey, macrophytes were present along all 27 transects with over 80% of transects having submerged macrophytes extending 60 m from shore. Of the quadrats sampled within 60 m of shore, nearly half had 40% or more cover from submerged macrophytes. A total of 6 species of submerged macrophytes were found (Potamogeton pectinatus, Elodea, Ceratophyllum, Vallisneria americana, Zannichellia palustris and Sagittaria) with Potamogeton pectinatus and Elodea being the most prevalent species.
E. Discussion
A short-term clear water phase was obtained following the third rotenone application involving elements of trophic cascading and changes in physical disturbance. One factor involved in determining water clarity in Ventura Marsh was suspended sediment as indicated by the negative correlation of inorganic suspended solids (ISS) and Secchi disk transparency (r2 = 0.51). Patterns in ISS and Secchi disk transparency tend to mirror one another (Fig. 3). Overall, ISS were not significantly lower during the clear water phase (BACI, p = 0.05), however, extremely low ISS values were recorded for approximately two months after the third fish removal. The low ISS values recorded in July and August of 2000 indicate a reduced amount of sediment in the water column. The benthic biomass during the clear water phase was approximately 5X greater than a similar time period during the turbid phase, suggesting that fish foraging was very low during the clear water phase. The high water clarity following the third fish removal may be partially due to lower amounts of suspended sediment as a consequence of reduced fish foraging. Similarly, Meijers et al. (1990) attributed increased water clarity in Lake Bleiswijkse Zoom and Lake Noorddiep following fish removal partially to decreased suspended sediment due to reduced bioturbation by fish.
Phosphorus concentrations appeared to be somewhat reduced during the clear water phase compared to the premanipulation period (BACI, p < 0.058). Sediment resuspension was low during portions of the clear water phase perhaps leading to reduced nutrient recycling and thus lower phosphorus concentrations. In addition, the reduced fish biomass may have led to lower total phosphorus concentrations due to reduced fish excretion.
The increased water clarity of Ventura Marsh appeared to also result from reduced phytoplankton biomass in the water column. Secchi disk transparency showed a strong negative relationship with phytoplankton biomass both in the turbid and clear water phases (r2 = 0.55, 0.74), indicating that water clarity increased as phytoplankton biomass declined. Phytoplankton biomass showed large fluctuations in 1999 but overall steadily increased throughout the summer (Fig. 6). A similar trend began to develop in 2000 with a ten-fold increase in phytoplankton biomass from March to early June. However, following fish removal in June phytoplankton biomass decreased to low biomasses, similar to those observed in 2000 following ice out, for approximately one month.
The reduction in phytoplankton biomass following the third fish removal appears to be due to zooplankton grazing. Shortly after the third fish removal, zooplankton grazing rates peaked at 55% of the marsh water per day, followed a month later by another peak of 139% of the marsh water per day (Fig. 9). The first peak was due solely to rotifers since cladocerans had not yet become established. During the second peak in grazing zooplankton biomass was not notably higher, but approximately half of the zooplankton biomass was comprised of Daphnia and Ceriodaphnia, which were rarely observed on other occasions (Fig. 7). Of the cladocerans and rotifers identified in this study, Daphnia and Ceriodaphnia were the genera with the highest filtration rates. Phytoplankton biomass remained low during these peaks in grazing suggesting that the control of phytoplankton was by zooplankton grazing.
The factor limiting phytoplankton biomass in Ventura Marsh switched between nutrients and zooplankton grazing. The amount of phytoplankton seemed to be determined by phosphorus concentration in the turbid phase (Fig. 4). Following the third fish removal Daphnia and Ceriodaphnia became abundant and phytoplankton was limited by zooplankton grazing. Zooplankton grazing rate quickly decreased from the high peak of 139% in July 2000 to less than 25% per day. The standing phytoplankton biomass may not have been sufficient to support the zooplankton community, thus resulting in a population decline. After the decline in grazing in late July 2000, phytoplankton biomass began to increase again. Grazing rate remained low
(≤35% per day) during August and September while phytoplankton continued to grow until constrained by another factor. The trend in standing phytoplankton biomass in August and September are similar to the trend in total phosphorus suggesting the phytoplankton was once again limited by phosphorus. Overall, phytoplankton in Ventura Marsh were phosphorus limited until after the third fish removal when it switched to top-down control. Two months later the system had reverted back to bottom-up control.
The maintenance of top-down control is essential to a successful fish biomanipulation. It is therefore important to discern the factors limiting the abundance of large-bodied filter feeding cladocerans. Top-down control occurred following the third fish manipulation for approximately two months when there was a substantial biomass of Daphnia and Ceriodaphnia. The reduction in fish predation and lower suspended sediment in the clear water phase may all account for the increase in biomass of these larger-bodied cladocerans. Biomass of Daphnia and Ceriodaphnia began to decline in late July, probably due to insufficient standing phytoplankton biomass. In August when the phytoplankton biomass began to increase again the cladoceran population did not respond with a subsequent increase in biomass and grazing rate, but remained low for the remainder of the study. An increase in juvenile carp and ISS, perhaps due to wind, may help explain why larger-bodied cladocernas did not become abundant again once their food source returned.
The
increase in water clarity was sufficient to promote a dramatic increase in
macrophyte diversity and abundance. The
higher water clarity and the reduced fish foraging allowed macrophytes to establish
at greater depths and in higher densities.
The presence of macrophytes is essential to the maintenance of high
water clarity. Macrophytes may help
maintain lower suspended sediment and biomass of phytoplankton (James and Barko
1990, Perrow et al. 1997, Van Donk et al. 1993).
F. Conclusions
The increased water clarity following the third fish removal can be partly explained by reduced physical disturbance and trophic cascading. A reduction in suspended sediment occurred due to lower fish foraging activity leading to increased water clarity. Reduction in phytoplankton biomass was also partially responsible for the increased water clarity. The low phytoplankton biomasses observed following the third fish manipulation can be explained by zooplankton grazing. The high grazing can be attributed mainly to the abundance of Daphnia and Ceriodaphnia. The reason for the increased abundance of these larger-bodied cladocerans is difficult to discern but likely contains elements of predation and turbidity.
Turbidity appears to be an important factor structuring Ventura Marsh. Water clarity, zooplankton composition, and macrophyte distribution may all be impacted by suspended sediment. A portion of the suspended sediment in the water column of Ventura Marsh can be attributed to the feeding activities of the benthivorous fish. Other factors, such as wind, may also contribute to the resuspension of sediment. The expansion of macrophyte beds following the clear water phase may help maintain lower turbidity in Ventura Marsh during subsequent years.
References
Adrian, Rita. 1991. Filtering and feeding rates of cyclopoid copepods feeding on phytoplankton. Hydrobiologia 210: 217-223.
American Public Health Association, American Water Works Association, and Water Environmental Federation. 1998. Standard Methods for the Examination of Water and Wastewater, 20th Edition. American Public Health Association, Washington, D.C.
Benke, A., A.D. Huryn, L.A. Smock, and J.B. Wallace. 1999. Length-mass relationships
for freshwater macroinvertebrates in North America with particular reference to
the southeastern United Phases. J. N. Am. Benthol. Soc. 18(3): 308-343.
Boers, P., L. Van Ballegooijen, and J. Uunk. 1991. Changes in phosphorus cycling in a shallow lake due to food web manipulations. Freshwat. Biol. 25: 9-20.
Bogdan, K.G., and J.J. Gilbert. 1982. Seasonal patterns of feeding by natural populations of Keratella, Polyartha, and Bosmina: Clearance rates, selectivities, and contributions to community grazing. Limnol. Oceanogr. 27(5): 918-934.
Bonneau, J.L. 1999. Ecology of a fish biomanipulation in a great plains reservoir. Dissertation. University of Idaho, Moscow, Idaho, USA.
Brooks, J.L., and S.I. Dodson. 1965. Predation, body size, and composition of plankton. Science 150: 28-35.
Brabrand, Å., B.A. Faafeng, and J.P.M. Nilssen. 1990. Relative importance of phosphorus supply to phytoplankton production: fish excretion versus external loading. Can. J. Fish. Aquat. Sci. 47: 364-372.
Breukelaar, A.W., E.H.R.R. Lammens, J.G.P.K. Breteler, and I. Tatrai. 1994. Effects of benthivorous bream (Abramis brama) and carp (Cyprinus carpio) on sediment
resuspension and concentrations of nutrients and chlorophyll a. Freshwat. Bio.
32: 113-121.
Brooks, J.L., and S. Dodson. 1965. Predation, body size, and composition of plankton.
Science 150: 28-35.
Carpenter, R.C., J.F. Kitchell, and J.R. Hodgson. 1985. Cascading trophic interactions and lake productivity. BioScience 35: 634-639.
Clarke, K.R., and R.M. Warwick. 1994. Changes in marine communities: an approach to statistical analysis and interpretation. Natural Environment Research Council, UK, 144pp.
Courtenay, W.R. Jr., D.A. Hensley, J.N. Taylor, and J.A. McCann. 1984. Distribution of exotic fishes in the continental United States. In W.R. Courtney Jr. and J.R. Stauffer Jr. (eds), Distribution, Biology, and Management of Exotic Fishes. John Hopkins University Press, Baltimore.
Crivelli, A.J. 1983. The destruction of aquatic vegetation by carp. Hydrobiologia 106: 37-41.
Crumpton, W.G., T.M. Isenhart, and P.D. Mitchell. 1992. Nitrate and organic N analyses with second-derivative spectroscopy. Limnol. Oceanogr. 37: 907-913.
Downing, J.A. 1979. Aggregation, transformation, and the design of benthos sampling programs. J. Fish. Res. Board Can. 36: 1454-1463.
Drenner, R.W., and K.D. Hambright. 1999. Review: Biomanipulation of fish assemblages as a lake restoration technique. Arch. Hydrobiol. 146: 129-165.
Driessen, O., B. Pex, and H.H. Tolkamp. 1993. Restoration of a lake: first results and problems. Verh. Internat. Verein. Limnol. 25: 617-620.
Dumont, H.J., I. Van de Velde, and S. Dumont. 1975. The dry weight estimate of biomass in a selection of Cladocera, Copepoda and Rotifera from the plankton, periphyton and benthos of continental waters. Oecologia 19: 75-97.
Eckblad, J.W. 1971. Weight-length regression models for three aquatic gastropod
populat ions. American Midland Naturalist 85(1): 271-274.
Findenegg, I. 1974. Expressions of populations. In R. A. Vollenweider (ed.), A Manual
on Methods for Measuring Primary Production in Aquatic Environments.
Blackwell Scientific Publications, Oxford.
Gosselain, V., C. Joaquim-Justo, L. Viroux, M. Mena, A. Metens, J.-P. Descy, and J.-P. Thome. 1996. Laboratory and in situ grazing rates of freshwater rotifers and their contribution to community grazing rates. Arch. Hydrobiol. Suppl. 113: 351-361.
Gulati, R.D. 1989. Structure and feeding activites of zooplankton communitiy in Lake Zwemlust, in the two years after biomanipulation. Hydrobiol. Bull. 23: 35-48.
Haney, J.F. 1973. An in situ examination of the grazing activities of natural zooplankton communities. Arch. Hydrobiol. 72: 87-132.
Hanson, M.A., and M.G. Butler. 1994. Response of plankton, turbidity, and macrophytes to biomanipulation in a shallow prairie lake. Can. J. Fish. Sci. 51: 1180-1188.
Hart, R.C. 1988. Zooplankton feeding rates in relation to suspended sediment content: potential influences on community structure in a turbid reservoir. Freshwat. Biol. 19: 123-139.
James, W.F., and J.W. Barko. 1990. Macrophyte influence on the zonation of sediment
accretion and composition in a north-temperate reservoir. Arch. Hydrobiol. 120:
129-142.
Kirk, K.L. 1991. Inorganic particles alter competition in grazing plankton: the role of selective feeding. Ecology 72(3): 915-923.
Lamarra, V.A. Jr. 1975. Digestive activities of carp as a major contributor to the nutrient loading of lakes. Verh. Internat. Verein. Limnol. 19: 2461-2468.
Lammens, E.H.R.R, R.D. Gulati, M.L. Meijer, and E. van Donk. 1990. The first biomanipulation conference: a synthesis. Hydrobiologia 200/201: 619-627.
Lammens, E.H.R.R., and W. Hoogenboezem. 1991. Diets and feeding behaviour. In I. J. Winfield & J. S. Nelsom (eds), Cyprinid Fishes Systematics, Biology and Exploitation. Chapman & Hall, London.
Lampert, W. 1987. Laboratory studies on zooplankton-cyanobacteris interactions. New Zealand Journal of Marine and Freshwater Research 21: 483-490.
Lougheed, V.L., B. Crosbie, and P. Chow-Fraser. 1998. Predictions on the effect of carp exclusion on water quality, zooplankton and submerged macrophytes in a Great Lakes wetland. Can. J. Fish. Aquat. Sci. 55: 1189-1197.
Lougheed, V.L., and P. Chow-Fraser. 1998. Factors that regulate the zooplankton community structure of a turbid, hypereutrophic Great Lakes wetland. Can. J. Fish. Aquat. Sci. 55: 150-161.
Lougheed, V.L. 2000. A study of water quality, zooplankton and macrophytes in wetlands in the Canadian Great Lakes basin: implications for the restoration of Cootes Paradise Marsh. Dissertation. McMaster University, Hamilton, Ontario, Canada.
Mason, W.T., and P.P. Yevich. 1967. The use of phloxine B and Rose Bengal stains to
facilitate sorting benthic samples. Trans. Amer. Microsc. Soc. 86: 221-223.
McCauley, E. 1984. The estimation of the abundance and biomass of zooplankton in samples. In J. A. Downing and F. H. Rigler (eds.), A Manual on Methods for the Assessment of Secondary Productivity in Fresh Waters. Blackwell Scientific Publications, Oxford.
Meijer, M.L., A.J.P. Raat, and R.W. Doef. 1989. Restoration of Lake Bleiswijkse Zoom
(The Netherlands): first results. Hydrobiol. Bull. 23: 49-57.
Meijer, M.L., M.W. de Haan, A.W. Breukelaar, and H. Buiteveld. 1990. Is reduction of
the benthivorous fish an important cause of high transparency following
biomanipulation in shallow lakes? Hydrobiologia 200/201: 303-315.
Meijer, M.L., and H. Hosper. 1997. Effects of biomanipulation in the large and shallow Lake Wolderwijd, The Netherlands. Hydrobiologia 342/343: 335-349.
Merritt, R.W., and K.W. Cummins. 1996. An Introduction to the Aquatic Insects of North
America. Third Edition. Kendall/Hunt Publishing, Dubuque, Iowa.
Mourelatos, S, and G. Lacroix. 1990. In situ filtering rates of Cladocera: Effect of body length, temperature, and food concentration. Limnol. Oceanogr. 35(5): 1101-1111.
Northcote, T.G. 1988. Fish in the structure and function of freshwater ecosystems: A “top-down” view. Can. J. Fish. Aquat. Sci. 45: 361-379.
Ozimek, T., R.D. Gulati, and E. van Donk. 1990. Can macrophytes be useful in biomanipulation of lakes? The Lake Zwemlust example. Hydrobiologia 200/201: 399-407.
Pennak, R.W. 1989. Fresh-water Invertebrates of the United Phases. John Wiley & Sons,
New York.
Perrow, M.R., M.L. Meijer, P. Dawidowicz, and H. Coop. 1997. Biomanipulation in
shallow lakes: phase of the art. Hydrobiologia 342/343: 355-365.
Post, J.R., and D.J. McQueen. 1987. The impact of planktivorous fish on the structure of a plankton community. Freshwat. Biol. 17(1): 79-90.
Reynolds, C.S. 1994. The ecological basis for the successful biomanipulation of aquatic
communities. Arch. Hydrobiol. 130: 1-33.
Roberts, J. A. Chick, L. Oswald, and P. Thompson. 1995. Effect of Carp, Cyprinus carpio L., an exotic benthivorous fish, on aquatic plants and water quality in experimental ponds. Mar. Freshwater Res. 46: 1171-1180.
Rojo, C., and J.
Rodríguez. 1994. Seasonal variability of phytoplankton size structure in a
hypertrophic lake. Journal of Plankton Research. 16(4) 317-335.
Rosen, R.A. 1981. Length-dry weight relationships of some freshwater zooplankton. J.
Freshwater Ecol. 1: 225-229.
Ruttner-Kolisko, A. 1977. Suggestions for biomass calculations of plankton rotifers.
Arch. Hydrobiol. Beih. Ergebn. Limnol. 8: 71-76.
Scheffer, M, 1998. Ecology of Shallow Lakes. Chapman & Hall, London. 357pp.
Schindler, D.W., and B. Noven. 1971. Vertical distribution and seasonal abundance of
zooplankton in two shallow lakes of the Experimental Lakes Area, Norhtwestern Ontario. J Fish. Res. Board Can. 28: 245-256.
Schindler, D.W. 1978. Factors regulating phytoplankton production and standing crop in the world’s freshwaters. Limnol. Oceanogr. 23(3): 478-486.
Schriver, P., J. Bøgestrand, E. Jeppesen, and M. Søndergaard. 1995. Impact of submerged macrophytes on fish-zooplankton-phytoplankton interactions: large-scale enclosure experiments in a shallow eutrophic lake. Freshwat. Biol. 33: 255-270.
Shapiro, J., V. Lamarra, and M.L. Lynch. 1975. Biomanipulation: An ecosystem approach to lake restoration. In P. L Brezonik & J. L. Fox (eds), The Proceedings of a Symposium on Water Quality Management Through Biological Control. University of Florida, Gainesville, Fl.
Shapiro, J and D. I. Wright. 1984. Lake restoration by biomanipualtion: Round Lake, Minnesota, the first two years. Freshwat. Biol. 14: 371-383.
Skubinna, J.P., T.G. Coon, and T.R. Batterson. 1995. Increased abundance and depth of submerged macrophytes in response to decreased turbidity in Saginaw Bay, Lake Huron. J. Great Lakes Res. 21(4): 476-488.
Smit, H., E.D. Van Heel, and S. Wiersma. 1993. Biomass as a tool in biomass
determination of Oligochaeta and Chironomidae. Freshwat. Biol. 29:37-46.
Smith, E.P., D.R. Orvos, and J. Cairns. 1993. Impact assessment using the before-after-
control-impact (BACI) model: concerns and comments. Can. J. Fish. Aquat. Sci.
50: 627-637.
Søndergaard, M, E. Jeppesen, P. Kristensen, and O. Sortkjær. 1990. Interactions between sediment and water in shallow and hypertrophic lake: a study on phytoplankton collapses in Lake Søbygård, Denmark. Hydrobiologia 191: 139-148.
Timms, R.M., and B. Moss. 1984. Prevention of growth of potentially dense phytoplankton populations by zooplankton grazing, in the presence of zooplanktivorous fish, in a shallow wetland ecosystem. Limnol. Oceanogr. 29(3): 472-486.
Van Donk, E., M.P. Grimm, R.D. Gulati, P.G. Heuts, W.A. de Kloet, and L. van Liere. 1990. First attempt to apply whole-lake food-web manipulation on a large scale in The Netherlands. Hydrobiologia 200/201: 291-301.
Van Donk, E., R.D. Gulati, A. Iedema, and J.T. Meulemans. 1993. Macrophyte-related shifts in the nitrogena nd phosphorus contents of the different trophic levels in a biomanipulated shallow lake. Hydrobiologia 251: 19-26.
Vanni, M.J., and D.L. Findlay. 1990. Trophic cascades and phytoplankton community structure. Ecology 71(3): 927-937.
Ward, H.B., and G.C. Whipple. 1959. Freshwater Biology. John Wiley & Sons, New
York.
Watson, S., E. McCauley, and J. A. Downing. 1992. sigmoid relationship between phosphorus, algal biomass, and algal community structure. Can. J. Fish. Aquat. Sci. 49: 2605-2610.
Welcomme, R.L. 1984. International transfers of inland fish species. In W.R. Courtney Jr. and J.R. Stauffer Jr. (eds), Distribution, Biology, and Management of Exotic Fishes. John Hopkins University Press, Baltimore.
Whitford, L.A., and G.J. Schumacher. 1984. A Manual of Freshwater Algae. Sparks Press Raleigh, N.C.
Wium-Anderson, S., U. Anthoni, C. Christophersen, and G. Houen. 1982. Alleopathic
effects on phytoplankton by substances isolated from aquatic macrophytes
(Charales). Okios. 39:187-190.
TABLE 1. Results of gill netting surveys on Ventura Marsh. Carp were categorized as either small (< 1.8 kg) or large (≥ 1.8 kg).
|
Date |
Common carp |
Size class |
Bullhead |
Channel catfish |
Premanipulation |
August 3, 1999 |
170 |
|
0 |
2 |
Postmanipulation 1 |
September 10, 1999 |
113 |
61 large 52 small |
31 |
1 |
Postmanipulation 2 |
April 12, 2000 |
84 |
9 large 75 small |
0 |
0 |
Postmanipulation 3 |
June 21, 2000 |
2 |
0 large 2 small |
0 |
0 |
Figure 1. Map of Clear Lake watershed showing land use and the location of Ventura Marsh and the reference system (Little Lake).
Figure 2. Map
of Ventura Marsh showing the water quality, plankton, and benthic sampling
sites
Figure 3. Inorganic suspended solids and Secchi disk transparency for Ventura Marsh from April 14, 1999 to September 27, 2000. Arrows indicate rotenone applications.
Figure 4 A & B. The relationship between total phosphorus and phytoplankton biomass was examined. Figure 4A is of the turbid phase (4/12/99 –6/6/00) and Figure 4B is of the clear water phase (6/7/00 – 9/27/00).
Figure 5. The relationship between inorganic suspended solids and Secchi disk transparency was examined. Solid circles indicate data from the turbid phase (4/12/99 –6/6/00) and open circle indicate data from the clear water phase (6/7/00 – 9/27/00).
Figure 6. Phytoplankton biomass and percent composition for Ventura Marsh from May 13, 1999 to September 27, 2000. Arrows indicate rotenone applications.
Figure 7. Zooplankton biomass and percent composition for Ventura Marsh from April 12, 1999 to September 27, 2000. Arrows indicate rotenone applications
Figure 8. Box-whisker plot of cladoceran and copepod lengths during premanipulation (4/12/99 – 8/16/99), postmanipulation 1 (8/17/99 – 10/15/99), postmanipulation 2 (3/14/99 – 6/6/00), and postmanipulation 3 (6/7/00 – 9/27/00). Copepod nauplii were not included in these plots.
Figure 9. Grazing rate and phytoplankton biomass for Ventura Marsh from May 13, 1999 to September 27, 2000. Arrows indicate rotenone applications.
Figure 10. Benthic biomass and percent composition for Ventura Marsh from April 12, 1999 to September 27, 2000. Arrows indicate rotenone applications
Figure 11. Box-whisker plot of chironomid and oligochaete lengths during premanipulation (4/12/99 – 8/16/99), postmanipulation 1 (8/17/99 – 10/15/99), postmanipulation 2 (3/14/99 – 6/6/00), and postmanipulation 3 (6/7/00 – 9/27/00).