The study was conducted at a marina site on Belton Lake (BL) (31.130394°N, -97.508359°W) and Stillhouse Hollow Lake (SHL) (31.020666°N, -97.525164°W) (Figure 1). Both lakes are U.S. Corps of Engineers reservoirs in Bell County, central TX, USA. They lie along the eastern edge of the Edward’s Plateau among limestone hills covered with juniper and small oaks. They are impoundments of the Leon and Lampasas Rivers, respectively, tributaries of the Little River which flows into the Brazos River (Figure 1). The sampling site at BL was on the south shore 1.12 km from its dam while that at SHL was on the lake’s north shore 1.90 km from its dam (Figure 1). It was the intent of this study to compare the population dynamics of zebra mussels at these two sites rather than at a number of different sites within each lake. A similar study of zebra mussel population dynamics at single sites among different water bodies including Belton Lake was conducted by Arterburn and McMahon (2022). Other comparative studies of zebra mussel population dynamics at single sites in different water bodies include those of Dorgelo (1993), and Chase and Bailey (1999).

BL (Figure 1) is located approximately 8 km northwest of the City of Belton, TX. Construction was completed in 1954 for flood control and conservation storage (Mendieta and Pate 1982). It serves as source water for local municipalities, hosts multiple recreational attractions, and is a popular site for recreational fishing and boating (TPWD 2022a). Seventeen public boat ramps occur throughout the lake. Its conservation pool elevation is 181 meters above mean sea level (m.a.s.l.), with a capacity of 0.53 km³ and encompasses 5012 ha (TPWD 2022a). Its mean and maximum depths are 11 m and 37.7 m, respectively (Tibbs and Baird 2019). Its shoreline length is 219 km of which 6% is made up of rock bluffs with traces of rocky shoreline and standing timber (Tibbs and Baird 2011). It has a watershed drainage area is 6630 km² (Holmquist et al. 2017a) and is considered mesotrophic with water clarity averaging 1.8 m (Tibbs and Baird 2019). In 2018, it had small amounts of aquatic plants but no noxious vegetation (Tibbs and Baird 2019). Zebra mussels were first reported in BL in 2013 (TPWD 2013).

SHL is approximately 8 km southwest of the City of Belton. Construction was completed in 1968 for flood control, recreation, and municipal water supply (Tibbs and Baird 2018). The lake has four public boat ramps. At conservation pool, its elevation is 190 m.a.s.l. with a capacity of 0.28 km³ and encompassing 2601 ha (Holmquist et al. 2017b; Tibbs and Baird 2018). It has mean and maximal depths of 11.0 m and 32.6 m, respectively, and 93 km of shoreline that is 39% rocky, 10% gravel, and 3% rock bluffs (Tibbs and Baird 2018). Its watershed is 3414 km² (TPWD 2022b). SHL is considered oligotrophic with sparse native aquatic plants. However, invasive Hydrilla occupied about half of the reservoir in 2017/2018 (Tibbs and Baird 2018). Zebra mussels were first reported in SHL in 2016 (TPWD 2016).

At each lake, water temperature and dissolved oxygen (DO) were recorded hourly at 1 m, 6 m, and 11 m depths from 02 September 2019 – 24 September 2020 with submersible Onset HOBO U26 data loggers (Onset Computer Corporation, Bourne, MA, USA). Data were downloaded monthly and daily mean values for temperature and DO calculated. The SHL logger at 6 m was lost during May 2020 resulting in data loss thereafter. Additionally, discrete monthly measurements of pH, temperature, and DO were taken at 1 m intervals from the water’s surface to the bottom substratum using a YSI ProODO Optical Dissolved Oxygen Instrument (YSI Incorporated, Yellow Springs, OH, USA). Its readings were used to confirm data logger outputs and to capture this data at 1-m intervals throughout the water column. Dissolved oxygen values were computed as a percentage of full air O₂ saturation (% DO) to allow O₂ concentration levels to be expressed independently of temperature. Surface water levels during the 2019–20 study at both lakes were obtained from TWDB (2021). Secchi depths were recorded during monthly site visits.

At each monthly site visit, five vertical plankton net (50 µm mesh, 20 cm aperture) tows from 10 m depth allowed estimation of veliger densities from October 2019 – January 2021. Plankton tow samples were transported to the laboratory on ice where they were filtered and preserved in 70% ethanol for later microscopic veliger enumeration. The five tows were estimated to draw 1570.8 l of water through the plankton net. While mixing a plankton sample to suspend veligers, a 5 ml sample was drawn from the sample jar and released into the channel of a clear Plexiglass Bogorov Plankton Counting Chamber in which veligers were counted at 35× power from one end of the chamber channel to the other with a dissecting microscope equipped with a polarized light attachment making veliger shells birefringent and easily identifiable (Johnson 1995). Separate veliger counts were made for each of 5 ml aliquots from each plankton sample giving a 25 ml sample volume. The total volume of the sample was determined by adding 25 ml to the remaining volume of the sample. The mean and standard deviation (sd) of veliger densities in the five, 5 ml sample aliquots were then used to estimate the total number of veligers and standard deviation (sd) in the original sample bottle by dividing the total original sample volume by 5 ml and multiplying that figure by the mean number of veligers in the 5, 5 ml counting sample aliquots. The standard deviations for mean number of veligers in the total sample were similarly estimated. The estimates of the total number veligers in the plankton sample and sd were then divided by the total volume of water passing through the plankton net (i.e., 1570.8 l) to estimate the number of veligers and sd per liter in the water column which was then multiplied by 1000 to yield an estimated mean number and sd of veligers in 1000 l of lake water.

To estimate mussel settlement and densities across depths and time, sixteen 20×20 cm PVC settlement plates were deployed bimonthly in each lake on 27 August 2019, 20 September 2019, 22 December 2019, 23 February 2020, 25 April 2020, and 29 June 2020. Each set of plates remained in place through 26 September 2020. The settlement plates were placed horizontally along a weighted nylon rope attached to the marina superstructure and extending vertically to just above the lake bed. Four settlement plates each with a total surface area of 0.085 m² separated by 5 cm PVC spacers were attached to the rope at depths of 1, 4.5, 8.4 and 11–12 m (sixteen plates total per rope). Readily identifiable mussels greater than approximately 2 mm in shell length (SL, the greatest linear distance from the anterior umbo to the posterior shell margin) on all deployed plates were counted monthly. Mussels with shell lengths <2 mm were not counted because they were too small to be accurately enumerated. Plates were exposed to air while shaded from sunlight only when mussels were being enumerated and were typically out of water for less than 10 minutes. To reduce mussel stress during counting, mussels were wetted with a spray bottle containing lake water. As in Locklin et al. (2020), at least two researchers enumerated mussels simultaneously to reduce the time plates were out of the water. After mussel enumeration, plates were re-deployed with mussels still attached.

Five settlement plates located at 6 m depth on a single rope were deployed in both lakes on 13 September 2019 to assess mussel shell length (SL) distributions and growth rates. Mussels were scraped from one of the plates on 23 May 2020, 26 June 2020, 25 July 2020, 29 August 2020, and 25 September 2020 with a sharp edged paint scraper and preserved in 70% ethanol. In the laboratory, sampled mussels were placed on a flatbed scanner and their shell lengths scanned at 600 dpi for measurement to the nearest 0.01 mm using Adobe Photoshop.

Lake water level variation throughout the sampling period was obtained from the Texas Water Development Board (TWDB 2021) and was similar in both lakes (Suppl. material 1). BL and SHL water levels declined to 0.93 and 1.15 m below conservation pool (CP), respectively, in February 2020. During spring rain runoff, BL and SHL water levels peaked at 0.80 and 1.00 m above CP in March 2020 followed by a subsequent decline to minimal levels of 0.88 in and 0.92 m below CP in September 2020.

Water temperatures at the two sampling sites were also similar. Daily means of hourly temperature data at 1 m (surface), 6 m (middle) and 11 m (bottom) showed little thermal stratification from sampling initiation in September 2019 through March 2020 (Suppl. material 2). At the 11 m depth, however, water temperatures were below those at 1 and 6 m beginning in March (SHL) and June (BL) indicative of thermal stratification during spring/summer. Mean daily maximum 1 m and 6 m depth water temperatures in both lakes were ≈30–31 °C at the beginning of sampling period in September 2019, fell to minimum values of ≈12 °C in February 2020 and returned to ≈30–31 °C in August 2020 (Suppl. material 2).

The pattern of water temperature variation discreetly measured monthly at 1 m intervals from the surface to a near bottom depth of 11 m (Suppl. material 3) was similar to that recorded hourly at 1 m, 6 m and 11 m depths in both lakes. Water temperatures were similar across all depths from September 2019 through the end of March 2020. Thereafter, thermal stratification was initiated and peaked on 26 July 2020 (BL: 28.9 °C at 1 m to 24.7 °C at 11 m; SHL: 28.9 °C at 1 m to 24.0 °C at 11 m). When last sampled on 25 September 2020, the two lakes were no longer stratified with water temperatures of 26.2 °C and 24.2 °C at 1 and 11 m, respectively. Paired T-tests indicated no significant differences (n = 14, p < 0.05) in water temperature between the lakes at any tested depth (Suppl. material 8).

Mean daily percent of full air O₂ saturation (%O₂) values were similar at both lake sites (Suppl. material 4). The daily mean %O₂ values were >60%O₂ approaching 100%O₂ from November 2019 to May 2020. Thereafter, at BL, %O₂ at 11 m declined to near 0% from July through September 2020 before again attaining near surface levels (>60%O₂). BL oxygen concentrations at 11 m during Fall 2019 as well as previous years (Locklin et al. 2020) resembled Fall 2020 concentrations suggesting acute hypoxia relatively deep in the water column is an annual event at BL. At 6 m, BL %O₂ values were intermediate between 1 m and 11 m from September through October 2019, becoming similar to surface levels from November 2019 through June 2020, and again being intermediate from June through September 2020 (Suppl. material 4: fig. S4A). Mean daily SHL %O₂ values at 1 m and 6 m depths were similar to those of BL from September 2019 through May 2020 after which the O₂ monitor at 6 m was lost. Like BL, at 11 m depth SHL %O₂ values were near 0% in September 2019, becoming roughly equivalent to those at 1 and 6 m (60–100%O₂) from October 2019 through March 2020. Thereafter, at 11 m, SHL, %O₂ progressively declined to <20%O₂ by the September 2020 cessation of sampling.

Discreet monthly %O₂ determinations at 1-m intervals also differed between the two sites with relatively little hypoxia occurring at any depth in SHL while BL experienced periods of hypoxia during September 2019 and from July through August of 2020 (Suppl. material 5). Paired T-tests indicated that BL %O₂ values were significantly lower at depths of 4–11 m than at SHL (Suppl. material 8), reflective of hypoxic events during July 2019 and June through August 2020 that did not occur at SHL (Suppl. materials 4, 5).

pH in both lakes was greatest from January through March and lowest from July through September (Suppl. material 6). Summer declines in pH were associated with corresponding periods of thermal stratification and decreased %O₂ in both lakes (Suppl. materials 3–6). Paired T-tests indicated that mean pH at BL across sampling dates was significantly lower than at SHL at all depths with the exception of 4 m (Suppl. material 8). The overall difference between mean pH values across all depths was 0.197 (sd ± 0.026), suggesting it was of little biological significance.

Secchi Disk depths recorded monthly from 29 September 2019 through 29 August 2020 were consistently higher at SHL than BL (Suppl. material 7). Over the sampling period, BL Secchi Disk values ranged from 1.2 m on 22 December 2019 to 2.67 m on 29 September 2019 while at SHL they ranged from 1.5 m on 19 January 2020 to 7.5 m on 24 April 2020. Mean Secchi disk depths were 2.04 m at BL and 3.96 m at SHL. A paired T-Test indicated that Secchi disk depths were significantly different between two lakes (p = 0.0011) suggesting that seston and, presumably, plankton concentrations were elevated in BL relative to SHL.

A seasonal pattern in mussel veliger density was recorded at both BL and SHL (Figure 2). In the initial 29 September 2019 sample, BL veliger densities were low at a mean of 4.3 (±9.5) veligers 1000 l^-1 when surface water temperature (SWT) was 27.8 °C. Veliger density decreased to 0 veligers 1000 l^-1 on 04 October 2019 when SWT = 24.7 °C (Figure 2; Suppl. material 2). Veliger densities at BL peaked at 129.5 (± 41.6) 1000 l^-1 on 27 October 2019 (SWT = 21.6 °C) and declined to 0 veligers 1000^-1 by 19 January 2020 (SWT = 12.8 °C). Veligers reappeared at BL on 23 February 2020 at a density of 1.0 (±4.7) 1000 l^-1 at a SWT of 11.4 °C. They remained at low densities (≤4.5 1000 1^-1) through 24 March 2020 (SWT = 15.7 °C), then peaked at 6,734.9 (±1,896.0) 1000 l^-1 on 23 May 2020 (SWT of 22.7 °C). BL veliger density then declined to a low of 19.7 (±13.7) 1000 1^-1 on 28 August 2020 (SWT = 28.2 °C). Thereafter, BL mean veliger densities rose to a peak of 1,698.3 (±213.4) 1000 l^-1 on 23 October 2020, thereafter declining to 23.6 (±8.8) 1000 l^-1 on the final sampling date of 1 January 2021 (Figure 2; Suppl. material 1).

Seasonal veliger density at SHL was similar to that of BL with the exception of not being sharply depressed on 28 August 2020 as recorded at BL. At SHL, no veligers occurred in the 20 September 2019 and 4 October 2019 samples (SWT = 29.1° and 28.3 °C). Veligers first appeared in the 27 October 2019 sample at 67.7 (±57.4) veligers 1000 l^-1 (SWT = 20.8 °C), following which densities peaked at 5,817.0 (± 851.8) 1000 l^-1 on 25 July 2020 (SWT = 28.9 °C). Thereafter, SHL veliger densities declined to 807.1 (±80.3) 1000 l^-1 through 25 September 2020 (SWT = 26.2 °C), and fell to a minimum level of 2.6 (±4.7) 1000 l^-1 on the final 01 January 2021 sample.

Mussels occurred on settlement plates after initial deployment on 02 September 2019 through final sampling on 26 September 2020 in both lakes (Figure 3). At the SHL site, densities were relatively stable from 27 October 2019 through 23 May 2020, thereafter increasing from 28 June 2020 to the final sample on 26 September 2020 (Figure 3G–L). At the BL site, mussels first settled on 27 October 2020 attaining maximum densities on 22 December 2019 (Figure 3A–F). Densities remained relatively stable through 23 May 2020 then increasing by 29 June 2020. Unlike the SHL site, BL site mussel densities on the 02 September 2019 deployed plates declined at depths of 9 m and below from 26 June 2020 through final sampling on 26 September 2020 (Figure 3A–F) associated with a concurrent decline in %O₂ from 26 June 2020 through 29 August 2020 ranging from 13%O₂ at 10 m depth on 29 August 2020 to 5.1%O₂ at 11 m depth on 26 August 2020 (Suppl. material 5: fig. S5A). At the SHL site, declines in %O₂ also occurred at 9–11 m depths but were less severe than at the BL site with a minimum of 36.3%O₂ at 11 m depth on 29 August 2020 (Suppl. material 5: fig. S5B). Similar fall declines in mussel density occurred on BL site settlement plates deployed on 22 December 2019 through 29 June 2020 (Figure 3C–F) during a period of hypoxia at depths ≥8 m (Suppl. material 5: fig. S5A).

Mussel settlement patterns were associated with veliger presence. Fall veliger presence at the BL and SHL sites (Figure 2) were followed by increasing settled mussel densities through 22 December 2019 (Figure 3) corresponding to peak veliger densities of 129.5 and 67.7 per 1000 l, respectively, on 27 October 2019. Thereafter, mussel densities remained relatively stable during a period of low veliger density from 24 November 2019 through 24 April 2020 (Figures 2, 3). Increasing veliger density beginning on 24 April 2020 (Figure 2) resulted in spring 2020 mussel settlement leading to maximum mussel densities on 28 August 2020 at both lakes (Figure 3).

When analyzed by a one tailed t-test assuming unequal variances, mean densities of mussels on the four settlement plates across depths of 1, 4.7, 8.4, and 11–12 m combined over the course of the study was significantly higher in SHL relative to BL (p range = 0.0030–0.049) at 10 of 11 sampling periods, and only not being significantly different (p = 0.139) in the 28 June 2020 sample (Figure 4). Across all four depths, mean mussel density over the entire sampling period was considerably higher at the SHL site (8291.4 ± 12475.6 m²) relative to the BL site (947.4 ± 1410.9 m²).

Long-term studies of zebra mussel population dynamics in three Texas water bodies, including the BL site, have indicated that they have spring and fall settlement cohorts that die out during the following year’s late summer-to-early fall period following chronic exposure to elevated water temperatures (Arterburn and McMahon 2022). Shell length/size-frequency distributions of mussels sampled from settlement plates deployed at a depth of 6 m at both sites on 13 September 2019 revealed presence of settled fall mussel cohorts (Figure 5) resulting from a fall 2019 mussel spawning period (Figure 2). The Fall 2019 cohort density at the BL site declined to 50 mussels m^-2 by the final 25 September 2020 sample (Figure 5A) and was no longer present at the SHL site after the 25 July 2020 sample (Figure 5B), suggesting that these cohorts had life spans of approximately one year.

Spring 2020 cohorts first settled on 26 June 2020 at both sites after appearance of high veliger densities on 23 May 2020 (Figure 2). These spring cohorts continued to increase in mean shell length (SL) through the final 25 September 2020 sample attaining significantly different mean SLs of 11.51 ± 3.32 mm and 8.48 mm ± 3.93 at the BL and SHL sites, respectively (Figure 5). Daily shell growth rates for spring cohorts were estimated by dividing their mean shell lengths on the final 25 September 2020 sample (11.51 ± 3.321 and 8.48 ± 3.93 mm for BL and SHL, respectively) by the 126 day growth period extending from 23 May 2020 to 25 September 2020, yielding estimated shell length growth rates of 91.4 and 67.3 µm/day for the BL and SHL Spring Cohorts, respectively.

The physiochemical similarity of BL and SHL sampling sites allows a relatively direct comparison of the impact of post-invasion time on their population dynamics. During our study, SHL veliger density was much higher than at the BL site. At the time of our study, zebra mussels were known to have been present in BL since 2013 and SHL since 2016. In 2015, two years after invasion, peak BL sampling site veliger density at 1–1.5 m depth was 165,054 1000 l ^-1. By 2017, peak densities had fallen to 19,281 1000 l^-1 (Arterburn and McMahon 2022), and further declined to 4,584 1000 l^-1 during this study, indicative of apparent population decline through time. At the SHL site, peak veliger density at 27,952 1000 l^-1 five years after invasion was similar to that at the BL site of 19,281 m^-2 in 2017 five years after invasion (Arterburn and McMahon 2022) indicative of potential density reductions through time at both sites. While not all zebra mussel populations experience long-term population declines (Strayer et al. 2019b), similar long-term zebra mussel population declines have been reported for other infested Texas water bodies including Lake Ray Roberts and Lake Texoma (Arterburn 2020; Arterburn and McMahon 2022) and other water bodies in North America (Strayer et al. 1996, 2019a, b; Petrie and Knapton 1999; Stangel and Shambaugh 2005; Strayer and Malcom 2006; NPS 2013; Hetherington et al. 2019; Rudstam and Gandino 2020) and Europe (Bij de Vaate 1991; Stańczykowska and Lewandowski 1992; Burla and Ribi 1998; Casagrandi et al. 2007).

Such long-term density declines after attaining density maxima as reported for zebra mussel populations in warm water southwestern US water bodies have been labeled as boom-bust dynamics by Strayer and Malcom (2006). That zebra mussels display boom-bust population dynamics in the three Texas lakes where mussel population dynamics have been examined, (i.e., Belton, Ray Roberts and Texoma) (this study, Arterburn 2020; Arterburn and McMahon 2022) suggests that it may occur in some other of the 33 Texas lakes now infested with zebra mussels. Although requiring further confirmation of long-term mussel population dynamic studies in southwestern US water bodies, long-term decline in some Texas zebra mussel populations could have important consequences for water body management, marinas and raw water using utilities and industries in terms of long-term efforts and financial investment for management of mussels in invaded water bodies and their macrofouling of water using facilities.

Mussel densities on settlement plates at the SHL site were consistently significantly higher than at the BL site corresponding to higher veliger concentrations during the spring 2020 reproductive period. At peak average density on 25 July 2020, SHL site density at 25,732 mussels m^-2 was 6.4 times greater than that of the BL site of 4,027 mussels m^-2. Earlier studies recorded much higher mussel densities at the BL site between 2015 and 2017 (Locklin et al. 2020; Arterburn and McMahon 2022) again potentially indicative of BL population decline.

Mussels on sampling plates at both sampling sites had life spans of approximately one year or less with individuals from the previous 2019 spring and fall reproductive periods dying out by July (SHL) or September (BL) during 2020. Similar late summer adult mussel die offs have been reported for other mussel populations in southwestern US water bodies including Oologah Lake, Oklahoma (Boeckman and Bidwell 2014), and Texas Lakes Belton (Locklin et al. 2020; Arterburn and McMahon 2022), Texoma and Ray Roberts (Arterburn and McMahon 2022) suggesting that it may be a relatively common mussel life history trait in southwestern US water bodies relative to the 3–5, up to nine year life spans reported for European populations (Karatayev et al. 2006) and the 1.5–4 year life spans reported for populations in the northern US and Canada (Chase and Bailey 1999).

The one year life spans of southwestern zebra mussel populations may be a result of adult mussels being unable to consume phytoplankton/bacterioplankton at rates needed to sustain metabolic demands leading to tissue loss (i.e., starvation) and eventual mortality as reported by Walz (1978), Dorgelo and Kraak (1993), Stoeckmann and Garton (1997) and Morse (2009). White et al. (2015) observed major adult zebra mussel die-offs when summer water temperatures exceeded 25 °C for long periods in Gull Lake, Michigan [i.e., 1700 degree hours (dh)] even though water temperatures remained below lethal levels. In this study from May 2020 through final sampling on 25 September 2020, zebra mussels at 1 and 11 m depths in BL site were exposed to 9,470 and 2,882 dh, respectively, and at SHL from 11,194 to 5,601 dh far exceeding that of 1,700 dh found lethal to adult Gull Lake mussels. Thus, elevated summer water temperatures in Southwestern US water bodies such as BL and SHL may induce mortality in the previous year’s settled mussels even though water temperatures rarely attain a lethal level of ≥32 °C for long periods (Boeckman and Bidwell 2014; Locklin et al. 2020; Arterburn and McMahon 2022).

At both sites, spring cohort mean shell lengths were significantly different on the final 25 September 2020 sampling date of 11.5 ± 3.3 (BL) and 8.5 ± 3.9 (SHL), allowing mean shell growth rates over a shell growth period of 126 days to be estimated as 91.4 and 67.3 µm/day, respectively. The elevated shell growth rates of BL relative to SHL mussels may have been a result of higher food availability and/or lower zebra mussel densities reducing food competition. Greater chlorophyll a concentrations at the BL site relative to the SHL site during 2021–2022 (Locklin, unpublished) are suggestive of increased food resources at BL. Overall shell growth rates for the 2020 spring mussel generations at both sites were lower than those reported in other southwestern US water bodies. During 2016–17 spring season mussel shell growth rates at the BL site were estimated to be 127.9 µm day^-1 (Locklin et al. 2020). Similarly, in Texoma Lake on the Texas/Oklahoma border, mean first season mussel shell growth rate was 121.4 µm day^-1 (Churchill et al. 2017), and in Lake Oologah, OK, 136.7 µm day^-1 (Boeckman and Bidwell 2014). In contrast, Arterburn and McMahon (2022) have reported spring cohort first season shell growth rates in three Texas water bodies that were similar to that which we recorded in BL. These included Belton Lake from 2015 to 2017 (mean = 87 µm day^-1, n = 3, range = 73–107 µm day^-1), Texoma Lake from 2011 to 2017 (mean = 84 µm day^-1, n = 7, range = 60–112 µm day^-1), and Ray Roberts Lake from 2014 to 2017 (mean = 89 µm day^-1, n = 4, range = 59–109 µm day^-1). Overall, first season zebra mussel shell growth rates in warm southwestern US water bodies are considerably elevated relative to populations in cooler, more northern North American and European water bodies (for a review see Locklin et al. 2020) (for zebra mussel first season shell growth rates at more northern latitudes see Clarke (1952), Stańczkowska (1964), Morton (1969), Bij de Vaate (1991), Dorgelo (1993), Dorgelo and Kraak (1993b), MacIsaac (1994), Mackie et al. (1995), Akcakaya and Backer (1998), Chase and Bailey (1999), Garton and Johnson (2000), Lewandowski (2001), Cope et al. (2006)). Rapid shell growth in warm southwestern US water bodies leads to early maturity and reproduction while elevated summer water temperatures result in approximately one-year mussel life spans (Boeckman and Bidwell 2014; Arterburn 2020; Locklin et al. 2020; Arterburn and McMahon 2022; this study). Rapid mussel growth rates in southwestern US water bodies appears to allow extensive mussel population expansion after initial invasion with attainment of maximum densities within five to six years of initial invasion. However, as indicated in this study and that of Arterburn and McMahon (2022), zebra mussels in Texas water bodies may display boom-bust population dynamics (Strayer et al. 2019a) with population density declining after initially attaining peak densities. However, further long-term studies of zebra mussel population dynamics in Texas and other southwestern US water bodies are required to determine if such boom-bust zebra mussel population dynamics are generally characteristic of mussels in warm southwestern US water bodies.