This study focused on Lake Austin and Lady Bird Lake in Austin, Texas, USA, that are the last two in-line impoundments of the Colorado River (Fig. 1). In Central Texas, USA, the Colorado River was impounded into a series of reservoirs for flood control, hydroelectric generation, and drinking water provisioning, but they are also popular recreational destinations.

Figure 1. 

Map of Lake Austin and Lady Bird Lake within Texas (inset; 30.26799, -97.74464). Locations of public boat ramps (black circles), and for collection of zebra mussel veligers (Emma Long Park and above Tom Miller and Longhorn Dams) and adults (black squares), and reservoir centerline and nearshore water temperatures and nutrients (gray circles), shown. Upper and lower delineations of Lake Austin at Emma Long Park and in Lady Bird Lake at Auditorium Shores. Refer to Table 1 for site number, name, and latitude and longitude.

Both reservoirs are constant-level, pass-through reservoirs, but Lake Austin experiences more motorboat traffic and is popular with wake surfing boats, water skiers, and fisherman. Lake Austin has four public and dozens of private ramps along its 32 km course. The most popular ramps are at the Pennybacker Bridge and Walsh Park in the lower reaches of the reservoir (Fig. 1, Table 1). The Lady Bird Lake reservoir is less than 10 km in length and has two public boat ramps, both located < 1 km from each other in the lower portion of the reservoir (Fig. 1, Table 1). Gas motorboats are banned, limiting watercraft to those using electric motors or trolling motors; instead, non-motorized recreational vessels (e.g., canoes, stand-up paddleboards) dominate the recreational activities. Evidence of a zebra mussel introduction into Lake Austin and Lady Bird Lake was discovered in 2017 and 2018, respectively (News Release: Feb. 15, 2018: TPWD Announces New Zebra Mussel Findings in Central Texas Lakes - TPWD, News Release: Aug. 17, 2017: Lake Austin Positive for Invasive Zebra Mussels - TPWD (texas.gov)).

The Lower Colorado River Authority (LCRA) sampled veligers approximately monthly from January 2018 to December of 2022 using a vertical tow net pull at two sites in Lake Austin and one site in Lady Bird Lake (Fig. 1, Table 1). The net was a 64 µm, 30.5 diameter x 91.4 cm length net with a weighted cod-end attached. The initial depth of the tow varied between 4 and 14 m depending on the depth of the epilimnion. Given the known diameter and length of the net, we were able to calculate the volume of water passing through as it was being pulled from depth. We repeated vertical tows until approximately 1,000 L of water were filtered. The cod-end was rinsed into a 1 L polyethylene bottle and fixed in the field with 200 proof non-denatured ethanol (EtOH), bringing the final sample solution to 50% EtOH. In the lab, preserved samples were buffered with NaHCO₃ at a rate of 0.1 g per 50 mL. We then allowed samples to settle in an Imhoff cone for 24 h, decanted down to 20 mL, and the final volume was transferred to 50 mL centrifuge tubes.

We vortexed the centrifuge tubes before taking a 1 mL aliquot and placing in a Sedgewick-Rafter cell counter; aliquot volume was adjusted to 0.25 mL if veliger counts exceeded 800. We conducted veliger counts of aliquots via cross-polarized light microscopy (Johnson 1995; Churchill and Quigley 2018). We determined final veliger concentration (# veligers L^-1) by repeating the counting process 5 × per sample, taking the mean of the five pseudo-replicates, multiplying by sample volume, and finally dividing by the volume of water filtered through the plankton net.

The CoA Watershed Protection Department (WPD) sampled adult zebra mussels at nine sites in Lake Austin and six sites In Lady Bird Lake using passive artificial substrates comprised of plates for population density estimates from 2018–2023 (Fig. 1, Table 1). To evaluate whether adult densities would initially be higher nearest boat ramps, we clustered sites closer, both upriver and downriver, to public boat ramps. New substrates were deployed, or existing substrates were cleaned of sediments and any attached mussels, in April prior to the anticipated spring spawn. Adult sampling occurred in November when substrate plates were scrapped clean, the zebra mussels were rinsed of sediment, and were then air dried in a greenhouse facility. We deployed two separate artificial substrates at each site to capture intra-site variability. Substrates were suspended from overhanging trees, docks, or herbivore exclosure pens in 0.5–2.5 m total water depth. The lack of suitable structures for suspending most substrates in the preferred 2–5 m water depth (Schwalb et al. 2023) meant they were kept in shallower water; however, we felt that this approach was valid as suitable natural substrates (e.g., riprap, tree limbs) were abundant in less than 2 m of water, and we observed that adults were abundant in the areas selected for monitoring.

We initially deployed artificial substrates comprised of Masonite plates (total surface area 0.44 m²; Science First; Yulee, FL, USA). However, having the artificial samplers in shallow water lead to multiple issues such as high flows flushing-out snags on which substrates were suspended or destroying the plate material, and excess sediment build up on horizontal plates. The latter two problems were addressed in 2022 with the deployment of new substrates of PVC (0.89 m²; Zebra Mussel Guy; Cedar Park, TX, USA) in both a horizontal and vertical position; the latter of which reduced sediment accumulation (Czarnołȩski et al. 2004). The new PVC substrates were deployed alongside the original Masonite substrates to ensure minimal substate or orientation bias. Additionally, there were repeated losses of substrates each year due to vandalism/removal despite signage. Due to the number of zebra mussels attached to the substrates at most sites, counting all the individuals was not feasible, so we created a standard curve relating the number of zebra mussel shells to shell dry weight (# zebra mussel = zebra mussel DW (g) / 0.1147; r² = 0.937).

We provided summaries of daily hydrologic discharges and water temperatures along with annual aquatic vegetation coverage as they varied significantly over the study period and are inter-related with changes in phytoplankton biomass and water quality (Bellinger et al. 2018). Water temperatures, partially influenced by flow volumes and water residence time, could also significantly impact the success of zebra mussels in a riverine reservoir system.

Hydrologic data was reported as daily average discharges (m³ s^-1) from the Mansfield and Tom Miller Dams and was provided by the LCRA. Surface water quality samples (0.3 m depth) were collected at three sites in the main stem of each reservoir by LCRA in Lake Austin and by the CoA WPD in Lady Bird Lake (Fig. 1, Table 1). Samples were collected approximately bimonthly (even months) from 2010–2016, after which monthly samples were generally collected monthly from spring through fall (May–October). Submerged aquatic vegetation extent and species composition surveys were carried out by the Texas Parks and Wildlife Department (see Bellinger and De Jesús 2024 for details).

We collected surface water temperature (°C) with a YSI Sonde (YSI Inc., OH, USA) during routine water quality monitoring and veliger sample collection at approximately 1 m depth. Nearshore water temperature data utilized in this study was collected as part of harmful algal proliferation monitoring program in Lake Austin and Lady Bird Lake. Discreet surface temperatures in 1–1.5 m of water was collected on a weekly to bi-weekly basis from June through October 2021–2023 in Lake Austin and 2020–2023 in Lady Bird Lake. Additionally, we used HOBO (Onset Brands, MA, USA) continuous temperature loggers on 1 h intervals in ~1 m of water for the months of May–October between 2021–2022 in Lake Austin and 2020–2022 in Lady Bird Lake.

The water quality parameters we reported here are known to be altered by the presence of zebra mussels and included Secchi disk depth (m), chlorophyll a (Chl a; µg L^-1; E445.0), total suspended solids (TSS; mg L^-1; SM2540D), ammonia (NH₃; µg L^-1; E350.1 NH3-N), nitrate/nitrite (NO_x-; µg L^-1; SM4500-NO3-H), total Kjheldal nitrogen (TKN; µg L^-1; E351.2 TKN), and total phosphorus (TP; µg L^-1; E365.4 Phosphorus). Total nitrogen (TN; µg L^-1) concentrations were determined as the sum of NO_x- and TKN. Samples for Chl a were collected into 250–1,000 mL amber plastic bottles; TSS samples were collected into 1,000 mL clear plastic bottles; NH₃, NO_X, and TKN were collected into 250 mL clear plastic bottles preserved with H₂SO₄^2-; and, water for TP was left unpreserved in a 250 mL clear plastic bottle. All samples were kept on ice and delivered either to the LCRA Environmental Services Laboratory (Austin, TX), or, after 2022, DHL Analytical Services (Round Rock, TX).

We averaged water quality parameters for the months of June–October to account for the months of peak phytoplankton biomass and zebra mussel growth and feeding. Phytoplankton biomass tends to be lowest in the winter months which would obfuscate the influence of zebra mussels.

We also grouped parameters into three time-periods based on the presence of submerged aquatic vegetation and zebra mussels in each reservoir. We felt it important to divide data into separate periods for analyses to account for previously documented changes in phytoplankton biomass in response to hydrology and SAV (Bellinger et al. 2018). In Lake Austin, the first data period (2010–2013) represented when SAV was present and zebra mussels were absent. Because of active management (Bellinger and De Jesús 2024), the second period represented three-years of no SAV, and we believe no appreciable biomass of zebra mussels (2014–2016). The final period from 2017–2023 represented when zebra mussels were abundant, there was a near total absence of SAV, and hydrologic flows had generally declined (except for a large flood event in late 2018 into 2019) due to an ongoing drought. In Lady Bird Lake, the period of SAV presence and zebra mussel absence was from 2010–2016. The next period captured zebra mussel colonization and a general lack of SAV in the reservoir (2017–2020). The final period saw an increase in SAV extent as flow rates have declined, coupled with zebra mussel presence (2021–2023).

Differences in water quality parameters between periods were analyzed with a one-way ANOVA and a Holm-Sidak post hoc analysis to test whether phytoplankton, TSS, or nutrients concentrations had changed with the presence of zebra mussels. For data that did not meet normality and equal variances after log₁₀ transformation, a Kruskal-Wallace test with Dunn’s post hoc comparison was used. A linear mixed-effects model was applied to evaluate the influence of SAV, previous 7 d average discharge, and adult zebra mussel densities on Chl a concentrations. Sites were modeled as random intercepts, and the model runs included a fixed effect for time. All continuous covariates were z-scored by subtracting the mean and dividing by the standard deviation. The R packages lme4 and lmerTest were used in the analyses (R Core Team 2021). All data are available through the Watershed Protection Department’s Water Quality Sampling Data Portal.

Veliger concentrations at Emma Long Park were lowest in 2018 (<10 veligers L^-1) and peaked early in 2020 at > 200 veligers L^-1 (Fig. 2A). Since 2020, spring veliger concentrations have generally remained above 30 L^-1. Conversely, above the Tom Miller Dam, veliger concentrations were highest in 2018 and 2019, and declined thereafter until a resurgence in 2022 (Fig. 2B). In Lady Bird Lake above the Longhorn Dam, veliger concentrations peaked between 2019 and 2020 (40–100 veligers L^-1) (Fig. 2C). Across all monitoring sites, veliger densities generally peaked between Julian days 120–175 when water temperatures were between 20–25 °C (Fig. 3). Veligers were observed through the summer and fall months with little spawning through the winter.

Figure 2. 

Temporal changes in zebra mussel veliger concentrations (# veligers L^-1) (bars; left y-axis) and water temperatures (°C) (solid line; right y-axis) at Lake Austin sites (A) Emma Long Park and (B) above Tom Miller Dam; and, (C) Lady Bird Lake above Longhorn Dam.

Figure 3. 

Changes in zebra mussel veliger concentrations (red circles, left y-axis) and water temperatures (°C) (black circles; right y-axis) based on Julian Day of the year. Veliger concentrations fit with a LOESS regression line and water temperatures fit with a peak Gaussian three parameter nonlinear regression line.

In Lake Austin the estimated peak average adult density occurred in 2019 at 3,560 zebra mussels m^-2 with a median of 2,922 zebra mussels m^-2 (407–5,160 zebra mussels m^-2) (25^th–75^th quartile) (Fig. 4A, Table 2). Thereafter, densities declined 86%, reaching a minimum average density of 475 zebra mussels m^-2 and a median of 30 zebra mussels m^-2 (7–169 m^-2) in 2022 (Fig. 4A). The large intra-annual variances were driven by longitudinal variability in abundances, with the lower reservoir having much greater densities from 2018–2020 and oscillating thereafter (Fig. 4B, Table 2).

Figure 4. 

Annual average adult zebra mussel densities (# m^-2) for (A) all sites in Lake Austin (closed circles) and Lady Bird Lake (open squares); and, (B) the upper half (closed circle and open box) and lower half (red circle and blue square) of Lake Austin and Lady Bird Lake, respectively.

In Lady Bird Lake average adult densities were estimated to have peaked in 2019 at 2,518 zebra mussels m^-2 with a median density of 1,764 zebra mussels m^-2 (850–3,531 zebra mussels m^-2) (Fig. 4A). Thereafter adult densities declined by three orders of magnitude, reaching a minima in 2023 system-wide of 2 zebra mussels m^-2 (Fig. 4A, Table 2). As in Lake Austin, average densities were highest at sites in the lower reservoir until 2022 when densities in the upper reservoir exceeded those of the lower reservoir (Fig. 4B, Table 2). Substrates were more frequently lost in Lady Bird Lake, notably downriver, reducing the number of sites with data but observations of natural substrates around monitoring locations and discussions with recreational vendors throughout the reservoir corroborated the declining presence of live adult zebra mussels throughout the reservoir over time.

Average daily discharges to Lake Austin (Fig. 5A) and Lady Bird Lake (Fig. 5B) were low during a drought of record between 2010 and 2015 relative to 2016–2020 when storage reservoirs were full. With the onset of another drought, discharges again declined after 2021.

Figure 5. 

Daily average discharges (m³ s^-1) from the (A) Mansfield Dam; and, (B) Tom Miller Dam. (C) Annual average submerged aquatic vegetation (SAV) extent (% reservoir area) in Lake Austin (solid bars) and Lady Bird Lake (open bars).

In Lake Austin, SAV extent expanded through 2012 coincident with low discharges, but was thereafter effectively eliminated largely due to the stocking of triploid grass carp (Bellinger and De Jesús 2024) and remained absent until 2023 (Fig. 5C). Lake Austin open water temperatures during veliger collection ranged from 12 °C to 31 °C (Fig. 6A). Coldest average temperatures (19.0 ± 1.0 °C) were observed right below the Mansfield Dam whereas highest temperatures were measured above the Tom Miller Dam (27.2 ± 0.8 °C). From the nearshore instantaneous observations, there were 32 measurements > 28 °C and 7 were > 30 °C (Fig. 6B). From the continuous (1 h interval) near-shore monitoring, 2021 was generally a warmer summer than 2022 (Fig. 6C). Between 2021 and 2022 there were 13 and 5, respectively, observation days with temperatures > 30 °C, but less than 12 observation hours > 32 °C.

Figure 6. 

Water temperatures (°C) collected from Lake Austin (left column) and Lady Bird Lake (right column). Top row was collected bi-monthly at 1 m depth in the reservoir thalweg; middle row were bi-weekly in the summer near-shore (< 1.5 m total depth); and, bottom row were collected hourly in the summer near-shore (<1.5 m total depth) (Table 1). The 30 °C horizontal line represents a commonly reported zebra mussel thermal tolerance threshold.

In Lady Bird Lake, SAV was abundant until the end of the drought period (i.e., 2015), after which elevated discharges and floods flushed out the vegetation (Fig. 5C). With the return of lower discharges beginning in 2020, SAV extent has again increased. Lady Bird Lake water temperatures were generally warmer than those of Lake Austin, with summer surface water temperatures averaging 26.6 ± 0.5 °C and varying by < 1.5 °C across the reservoir (Fig. 6D). From the bi-weekly nearshore measurements, there were a total of 53 observations > 28 °C, 10 measurements > 30 °C, and two observations > 32 °C, both at downriver site 1997 (Fig. 6E). As in Lake Austin, 2021 had the most hourly measurements with water temperatures over 28 °C and 30 °C (Fig. 6F). Across sites in 2021, there were 60–80 observation days with temperatures > 28 °C, and 9–21 days with temperatures > 30 °C. Water temperatures > 32 °C were experienced most frequently at downriver site 1997 (Fig. 6F).

Lake Austin summer Chl a and TSS concentrations (ANOVA f_2,97 = 15.9, p < 0.001 and Kruskal-Wallis H = 24.7, df, 2, p < 0.001, respectively), and Secchi disc depth (Kruskal-Wallis H = 23.0, df = 2, p < 0.001) significantly differed between periods (Fig. 7A–C, Table 3). Measures were similar in the period of SAV presence (2010–2013) and after zebra mussel colonization (2017–2023).

Figure 7. 

Lake Austin (left column; closed circles) and Lady Bird Lake (right column; open circles) temporal patterns of water quality parameters: top row Chlorophyll a (Chl a; μg L^-1); middle row total suspended solids (TSS; mg L^-1); and, bottom row Secchi disc depth (m). Data were fit with LOESS regression. Lake Austin periods represent presence of submerged aquatic vegetation (SAV) and no zebra mussels (ZM) (2010–2013), absence of both SAV and ZM (2014–2016), and absence of SAV presence of ZM (2017–2023). Lady Bird Lake periods represent presence of SAV and absence of ZM (2010–2016), absence of SAV and presence of ZM (2017–2020), and presence of both SAV and ZM (2021–2023).

Lake Austin surface NH₃ concentrations were significantly different between periods (Kruskal-Wallis H = 10.3, df = 2, p < 0.01), whereas NO_x- (Kruskal-Wallis H = 2.4, df = 2, p = 0.31) and TN (Kruskal-Wallis H = 5.7, df = 2, p = 0.06) did not significantly differ (Fig. 8A–C, Table 3). Surface water TP concentrations were significantly higher after 2017 compared to the previous periods (Kruskal-Wallis H = 11.6, df = 2, p < 0.01; Fig. 8D). With the significant increase in TP concentrations, molar N:P ratios after 2017 significantly declined (Kruskal-Wallis H = 16.1, df = 2, p < 0.001; Fig. 8E; Table 3). Bottom nutrients from site 573 significantly differed in concentrations of NH₃ (Kruskal-Wallis H = 9.8, df = 2, p < 0.01) and TP (Kruskal-Wallis H = 8.9, df = 2, p < 0.05), and N:P ratios (Kruskal-Wallis H = 31.8, df = 2, p < 0.001) after 2017 relative to prior periods. Conversely, NO_x- and TN concentrations did not significantly differ between periods (ANOVA F < 1.5, df = 2, P > 0.05 and Kurskal-Wallis H < 8.8, df = 2, p > 0.05; Fig. 8A–E, Table 4).

Figure 8. 

Lake Austin (left column) and Lady Bird Lake (right column) temporal patterns of nutrients from the surface (0.3 m; blue circles) and bottom (0.5 m above bottom; red squares) for ammonia (NH₃; μg L^-1 ; top row); nitrate+ nitrite (NOx-; μg L^-1; second row); total nitrogen (TN; μg L^-1; third row); total phosphorus (TP; μg L^-1; fourth row); and molar N:P (fifth row). Data were fit with LOESS regression with solid lines for surface and dashed lines with bottom nutrients. Lake Austin periods represent presence of submerged aquatic vegetation (SAV) and no zebra mussels (ZM) (2010–2013), absence of both SAV and ZM (2014–2016), and absence of SAV presence of ZM (2017–2023). Lady Bird Lake periods represent presence of SAV and absence of ZM (2010–2016), absence of SAV and presence of ZM (2017–2020), and presence of both SAV and ZM (2021–2023). Note y-scale differences between reservoirs.

In Lady Bird Lake from 2010–2016 Chl a (Kruskal-Wallis H = 5.7, df = 2, p = 0.06) and TSS concentrations (Dunn’s Q > 2.5, p < 0.05) were higher, and Secchi depth significantly lower (Dunn’s Q > 2.9, p < 0.05), than the period of SAV absence and zebra mussel presence (2017–2020) (Fig. 7D–F, Table 3). After 2021, zebra mussel populations declined and SAV spread throughout the reservoir. During this period, Chl a and TSS concentrations increased relative to the prior period but were still lower than the 2010–2016 period of SAV presence/zebra mussel absence.

Lady Bird surface concentrations of NH₃ (Kruskal-Wallis H = 0.6, df = 2, p = 0.74) and TN (ANOVA F_2,115 = 1.3, p = 0.27) did not change significantly between periods whereas NO_x- concentrations were significantly greater during the 2017–2020 period (Kruskal-Wallis H = 8.0, df = 2, p < 0.05; Fig. 8F–H, Table 3). Concentration of TP have been highest (Kruskal-Wallis H = 17.2, df = 2, p < 0.001) and N:P lowest (Kruskal-Wallis H = 21.4, df = 2, p < 0.001) after 2021 (Fig. 8I, J, Table 3). Bottom water chemistry from sites 2 and 5 differed between periods for TP concentrations and molar N:P ratios (Holm-Sidak t > 2.5, p < 0.05), but not for NH₃, NO_x-, and TN (Kruskal-Wallis H = 0.6, df = 2, p = 0.74; Fig. 8F–J, Table 4).

The linear mixed-effects models indicated that estimated fall adult zebra mussel densities did not have a meaningful impact on the Chl a concentrations measured over the previous months (p > 0.05). However, the 7-d average discharges prior to sampling of Chl a was a significant predictor (p < 0.05).

Lake Austin zebra mussel veliger concentration estimates (>100 L^-1) were among the highest documented in Texas (Churchill et al. 2017; Hallidayschult et al. 2021; Locklin et al. 2024), though veliger estimates may be inflated due to accumulations in the water column near the dams. Adult population estimates across sites and years of up to ~10,000 m^-2 were similar to some systems (James et al. 2000; Alix et al. 2016), but much lower than observed elsewhere (>40,000 m^-2; Idrisi et al. 2001; Karatayev et al. 2015), including central Texas (>80,000 m^-2; Locklin et al. 2024). Estimates from Austin’s reservoirs may be on the low side of the actual population density due to substrate losses or the shallow depth of substrates. Our adult population estimates were initially highest in the lower reservoir and subsequently increased in the central and upper sections. The slow spread of zebra mussels up-river may have been due to overall lower recreational transportation of viable mussels or because of lower inputs of viable veligers from Lake Travis through the Mansfield Dam. The Mansfield dam is a deep hypolimnetic discharging dam (>30 m), though the reservoir generally remains well mixed in the spring during the initial zebra mussel spawn possibly enabling a small number of viable veligers to be introduced downriver. Additionally, as water levels have declined since 2018 (as of 28 April 2023 Lake Travis was ~ 12 m below full), the potential for discharges of veligers to Lake Austin would be expected to be greater.

The consistently cool discharges through the Mansfield Dam to Lake Austin did help keep water temperatures throughout most of the reservoir below 30 °C. In sub-tropical climates, it appears that heat-adapted zebra mussels require prolonged exposure to temperatures above 30 °C before veliger or adult mortality occurs (Spindle et al. 1995; Allen et al. 1999; Schwalb et al. 2023). It was only at the most downriver site that we observed water temperatures that could be considered stressful, notably after 2021 when flows were lowest. Additionally, Chl a concentrations averaged below 5 μg L^-1 after colonization which may have contributed to resource limitations, resulting in lower survival of juvenile and adult zebra mussels (Jantz and Neumann 1998; Palais et al. 2011; Churchill et al. 2017). With warmer water temperatures, zebra mussels have greater metabolic demands and therefore require more resources than might be available (McMahon 1996; Arterburn and McMahon 2022). Variable phytoplankton biomass condition observed after zebra mussel colonization was likely exacerbated by the increased daily flow volumes between 2016–2019, flushing algal biomass from the reservoir (Bellinger et al. 2018).

Along with the changes in TSS and phytoplankton concentrations, we also observed significant differences in water TP and NH₃ concentrations, and molar N:P ratios after zebra mussel colonization. Regeneration of P as soluble reactive phosphorus and ammonium by zebra mussels have been well documented (James et al. 1997, 2000; Strayer et al. 1999). Given that SRP concentrations were below lab detection limits (i.e., ~8 μg L^-1), and phytoplankton and TSS concentrations were reduced, we can only hypothesize that the increase in TP concentrations was driven by bacteria and picoplankton not captured on a glass fiber filter (Findlay et al. 1998; but see Carrick et al. 2015). The observed elevated NH₃ concentrations would be due to the pH of Lake Austin (i.e., > 8) shifting the balance from ammonium (NH₄⁺), though it is unclear why nitrification did not convert regenerated NH₃ to NO_x- in the oxygen-rich waters (Reddy and DeLaune 2008). Zebra mussel filtering of phytoplankton and suspended solids also typically results in benthification of nutrients (Mayer et al. 2003; Karatayev et al. 2015; Ozersky et al. 2015), but documenting this ecosystem change in Lake Austin from the long-term water quality monitoring sites was complicated. For example, site 560 is too shallow and strongly influenced by the Mansfield Dam discharges whereas site 561 is too deep and stratifies in summer. At site 573 there were no clear trends in bottom water quality changes after zebra mussel colonization, suggesting zebra mussels had an overall negligible impact. Supporting the apparent lack of widespread benthification of Lake Austin was an absence of Cladophora mats as is typical in systems after colonization by zebra mussels (Francoeur et al. 2017).

Lady Bird Lake veliger and adult population estimates also peaked in 2019 and 2020, soon after colonization, but veliger and adult estimates were lower than those of Lake Austin. Longitudinally, adult densities were highest in the lower reservoir, nearest the two boat ramps, suggesting they were the point of introduction. Lake Austin stratifies later than Lake Travis, providing opportunity for a small number of veligers to mix and be discharged to Lady Bird Lake, which may account for the delayed upriver increase observed. However, unlike Lake Austin, the Lady Bird Lake population experienced a dramatic decline (~3 orders of magnitude) after 2021. Similar, rapid declines in zebra mussel densities have been documented in the Hudson River post colonization (Pace et al. 2010). We did observe an increase in veliger concentrations in 2022 but documented adult densities remained low in 2023, suggesting that the zebra mussel population may be more robust in deeper water than nearshore where we were monitoring. We do not believe declines in adult densities were due to resource limitations as Chl a concentration generally remained above 5 µg L^-1 throughout the summer months. Therefore it seems probable that persistently elevated summer water temperatures contributed to the nearshore declines in zebra mussel densities.

Water temperatures in the 9.7 km-long Lady Bird Lake were generally warmer and varied much less longitudinally than in the 33.8 km long Lake Austin, despite the Tom Miller Dam also having a hypolimnetic discharge. Discharge volumes from Tom Miller Dam were lower than from the Mansfield Dam, and the latter dam discharges from a deeper, colder hypolimnion. While water temperatures taken from the centerline of Lady Bird Lake found few occurrences over 30 °C, it is likely the lack of consistently cool water (i.e., < 20 °C) inputs contributed to the observed nearshore water temperatures that were more frequently above thresholds likely to elicit thermal stress in veligers and adults. For example, in 2021 decreased discharges from the Tom Miller Dam coincided with week-long periods of water temperatures > 30 °C. Even for those zebra mussel populations better adapted to the southern United States, mass mortality events can occur during prolonged exposure to temperatures at and above 28 °C (Spindle et al. 1995; Allen et al. 1999; White et al. 2015; Churchill et al. 2017).

The period of SAV absence, increased flows, and abundant zebra mussels (i.e., 2017–2021) resulted in lowest Chl a concentrations and elevated water clarity heretofore not seen in the summer months. However, when all data years were considered, it was only the changes in discharge volumes that were significantly related to phytoplankton biomass. While it is not surprising that discharge played an important role in the amount of Chl a measured (Bellinger et al. 2018), that zebra mussels were found not to be a significant contributor suggests that in pass-through reservoirs, zebra mussel densities may play a smaller role in regulating phytoplankton biomass than discharge dynamics. With the recent declines in average daily flows through Lady Bird Lake, SAV extent has rapidly increased. Contrary to expectations, the increased biomass of SAV, coupled with zebra mussel presence, has not negatively affected phytoplankton biomass. Instead, clarity has decreased coincident with increasing TSS and Chl a concentration. Both measures were also significantly higher during the pre-zebra mussel period when SAV was abundant, suggesting multiple stable states (i.e., clear-water SAV and turbid phytoplankton-dominated) can co-occur in Lady Bird Lake based on the reservoir’s morphology (Ibelings et al. 2007; Janssen et al. 2014).

We observed significant changes in surface and bottom water nutrients between periods. Surface water NO_x- was highest during the period of ZM presence and SAV absence. Greater NO_x- concentrations have been documented post-zebra mussel colonization driven by oxidation of excreted NH₃/ NH₄⁺ (James et al. 1997; Higgins et al. 2008). Additionally, the pH of Lady Bird Lake is lower (i.e., < 8) than Lake Austin which favors NH₄⁺ over NH₃, but the former was not measured as part of the long-term monitoring program. Surface and bottom water TP concentrations have been increasing since 2017, after zebra mussel colonization, but were highest after 2021. We also observed an increase in benthic Cladophora mats after 2017, though mats now co-occur with Cabomba caroliniana and Utricularia spp., primarily in the upper reservoir which is shallower, has higher water clarity, and a substrate more conducive to zebra mussel colonization (i.e., cobble and boulders). The changes in bottom water nutrients are therefore likely being influenced not just by zebra mussel translocation of materials to the benthos, but also senescence of the abundant algae and SAV which would reduce oxygen concentrations at the sediment/water interface altering redox conditions (Reddy and DeLaune 2008). It is also likely that the use of lanthanum-modified bentonite throughout the summer months around Red Bud Isle since 2021 is altering sediment P-release dynamics around that long-term monitoring location (Copetti et al. 2016).