A recent Dreissenid molecular phylogeny using COI and 16S sequences (Therriault et al. 2004) concluded that Dreissena bugensis and D. rostriformis represent distinct races of a single species (sequence divergence was 0.23-0.54%, comparable to intraspecific variability in other species). The authors recommended the ancestral name D. rostriformis be used, with two races, bugensis and distincta (Therriault et al. 2004, Orlova et al. 2005). Similar suggestion was recently made by Karatayev and Burlakova (2022).

European regional assessment: Least Concern (LC)
EU 27 regional assessment: Not Applicable (NA)

The Quagga Mussel (Dreissena rostriformis ssp. bugensis) is assessed as Least Concern for the European region due to its moderately widespread, abundant native distribution. It is tolerant of a variety of habitats and demonstrates tolerance to habitat alteration. It does not occur as native in the EU27 Member States and is therefore assessed as Not Applicable for the EU27 Member States.

The species has highly invasive tendencies and it is a major invasive species in European and North American freshwater systems and is threatening native fauna in these regions.

Currently, the native range of D. r. bugensis is thought to include the entire Dnieper-Bug Liman system (a large coastal lake connected to the Black Sea estuary), the Dnieper River Delta, and the lower reaches of the South Bug and Ingulets rivers in Ukraine (Son 2007, Zhulidov et al. 2010). In 1941, D. r. bugensis was introduced into the Dnieprovskoe reservoir (Ukraine). The species is native-endemic to the European region.

With the construction of new reservoirs on the Dnieper River in the 1950s and 1960s, as well as canals for commercial shipping and irrigation, D. r. bugensis began to spread throughout Eastern and Western Europe. During the first four decades of the invasion (1940s-1980s), the spread of D. r. bugensis was slow, but increased substantially in the 1990s (reviewed in Orlova et al. 2004, Karatayev et al. 2011, Karatayev and Burlakova 2022a). In 2003, D. r. bugensis was found in the Moscow River within the city of Moscow (Lvova 2004), in 2006 the species was first recorded in the Netherlands (Molloy et al. 2007) and then rapidly spread to other parts of Western Europe. By 2021, D. r. bugensis was present in 16 European countries and four additional geographical provinces within large countries (Karatayev and Burlakova 2022a). In 1989, D. r. bugensis was found in North America, and in 2023, D. r. bugensis was recorded in 19 states and two provinces of three North American countries (Canada, Mexico and the USA) (Benson et al. 2023, Karatayev and Burlakova 2022a).

The main pathways of range expansion for this species include downstream drift of mussel larvae, attachment of mussels to boats, and ballast water discharge (Karatayev and Burlakova 2022a). Major geographic expansions have been associated with human activities that have provided previously unavailable means of dispersal, such as overland transport of boats and fishing gear, construction of reservoirs for water storage and power generation, construction of shipping channels for trade, changes in the nature and volume of international trade, and newer industrial practices and environmental regulations (Karatayev et al. 2007).

This species, together with the congener D. polymorpha, dominates/co-dominates benthic communities in its native range in the shallow eutrophic Dnieper-Bug Liman estuary in Ukraine (Markovskiy 1954). Populations of D. r. bugensis in the native range appear to be stable (Markovskiy 1954; L. Shevchuk pers. obs.).

In the invasive range, D. r. bugensis has steadily increased in abundance since the 1980s and 1990s in both Europe and North America (reviewed in Karatayev et al. 2011; bij de Vaate et al. 2014, Karatayev and Burlakova 2022a). The introduction of D. r. bugensis into lakes and reservoirs previously colonised by D. polymorpha often leads to dramatic declines in D. polymorpha populations (Orlova et al. 2004, 2005; Balogh et al. 2018, Karatayev et al. 2011, 2015, 2021a; bij de Vaate et al. 2014, Strayer et al. 2019). However, the outcome of this competition varies from almost complete extirpation of D. polymorpha to coexistence of both species and sometimes even reversal to D. polymorpha dominance (Zhulidov et al. 2010; Karatayev et al. 2011, 2021b). The outcome of competition largely depends on lake morphometry, dominant substrate types, food availability, predation, etc. (reviewed in Karatayev and Burlakova, 2022a and references therein). In general, in shallow polymictic lakes, D. r. bugensis becomes dominant 4-12 years after coexisting with previously established D. polymorpha, but does not appear to completely replace D. polymorpha. In large, stratified, deep lakes, D. r. bugensis becomes dominant more quickly, reaching much higher densities and driving D. polymorpha to virtual extinction (Karatayev et al. 2021a, Karatayev and Burlakova 2022b).

Habitats
This species occurs in a range of habitats from freshwater to oligohaline (up to 3-4 parts per thousand) in lakes, reservoirs, ponds, estuaries and brackish coastal lakes, but the highest densities of D. r. bugensis occur in deep stratified lakes (Karatayev et al. 2021a, Karatayev and Burlakova 2022a, 2022b). In shallow areas exposed to wave action, D. r. bugensis requires a hard surface for attachment, such as rocks, shells and coarse sand, but in deep stratified lakes D. r. bugensis forms extremely high densities on soft substrate, often covering >80% of the bottom in profundal areas (Karatayev et al. 2018a, Karatayev and Burlakova, 2022b). D. r. bugensis also forms high densities on various artificial substrates (Karatayev and Burlakova 2022a). Environmental factors controlling D. r. bugensis include substrate and food, eutrophication, pollution, oxygen depletion, competition and predation (reviewed in Karatayev and Burlakova 2022a). In the central basin of Lake Erie at depths >20 m, where bottom hypoxia routinely develops by the end of the growing season, only small numbers of juvenile mussels are found, indicating that hypoxia has eliminated the adult mussels that settled the previous season and limited recruitment and survival of juvenile D. r. bugensis (Karatayev et al. 2018b, 2021b). Therefore, monitoring the distribution of Dreissena can be an effective tool for mapping the extent and frequency of hypoxia in freshwaters. The upper temperature limit of D. r. bugensis observed in the field is 30-31°C (Karatayev et al. 1998, 2007, 2014; Wong et al. 2012; Karatayev and Burlakova 2022a). In brackish waters, D. r. bugensis thrives in the least saline areas, from freshwater and up to 3-4‰ (Mills et al. 1996, Orlova et al. 2005, Orlova 2014; Karatayev et al. 2007, 2014).

Ecological impacts
Both Dreissena species (D. polymorpha and D. r. bugensis) are sessile suspension feeders that attach to the substrate with byssal threads and form three-dimensional reef-like structures that alter the benthic environment (local impacts). They also affect the planktonic community, trophic relationships and nutrient cycling through their feeding activity (system-wide impacts). Virtually all studies of the local ecosystem effects of Quagga Mussels have been conducted in lakes that were first invaded by Zebra Mussels and then by Quagga Mussels, making it very difficult to disentangle the effects of the two species (Karatayev and Burlakova 2022a).

Local impacts
Similar to D. polymorpha, D. r. bugensis is an ecosystem engineer that can physically modify bottom substrates, creating new three-dimensional structures that provide shelter for benthic organisms from predation and other stressors (waves, currents, desiccation) that would otherwise be absent in this environment. Therefore, both species may have beneficial effects on native littoral epifaunal invertebrates (e.g. amphipods, isopods, leeches, turbellarians, hydrozoans, and some oligochaetes and chironomids). To date, much more information is available on D. polymorpha than on D. r. bugensis, but the limited data available suggest that the impact of both species in the littoral zone may be similar (reviewed in Karatayev et al. 2015, Karatayev and Burlakova 2022a). In the profundal zone of deep stratified lakes, such as the North American Laurentian Great Lakes, D. r. bugensis does not form large multilayered beds, often covering the entire bottom in a single layer, and its impacts on other benthic invertebrates are less pronounced, although positive effects on chironomids and oligochaetes have been reported (Karatayev and Burlakova 2022b, Eifert et al. 2023). In contrast, native filter feeders such as sphaeriids and the deep-water detritivorous amphipod Diporeia sp. In the Great Lakes, D. r. bugensis is negatively affected due to trophic competition in this food-limited profundal (reviewed in Karatayev et al. 2015; Karatayev and Burlakova 2022a, 2022b).

System-wide impacts
As both Dreissena species are suspension feeders, they may have similar ecosystem impacts on the water bodies they invade, suggesting that the magnitude of their impacts should be directly related to their biomass and that the impacts should be additive. However, the ability of Quagga Mussels to survive and grow in lower food and temperature regimes than Zebra Mussels allows them to consume phytoplankton, reduce chlorophyll concentrations and increase Secchi depths more effectively than Zebra Mussels, both earlier and later in the growing season, resulting in greater ecosystem impacts per unit of mussel biomass (Karatayev et al. 2002). The expansion of Quagga Mussels into deeper areas of the dimictic Lakes Ontario, Michigan, Huron and the eastern basin of Lake Erie in the 2000s was associated with almost complete extirpation of Zebra Mussels and a dramatic increase (22- to 45-fold) in lake-wide Quagga Mussel biomass (Karatayev et al. 2011; Karatayev and Burlakova 2021a, 2022b), causing strong and consistent system-wide changes, including increases in Secchi depth and decreases in total phosphorus concentrations, chlorophyll a, phytoplankton primary production, and phytoplankton and zooplankton biomass. The phytoplankton declines resulted from the almost complete disappearance of winter and spring diatom blooms, which in turn increased silica concentrations in the water column (reviewed in Karatayev and Burlakova 2022a, 2022b). The significant decrease in total phosphorus concentrations associated with dreissenids led to oligotrophication of all deep Great Lakes (reviewed in Karatayev and Burlakova 2022a, 2022b). Quagga Mussels are now the primary driver of phosphorus cycling in the four lower Great Lakes, providing a dramatic example of a large-scale reorganization of biogeochemical cycling caused by a single organism (Li et al. 2021).

Impacts on unionids
Heavy shell infestation by dreissenids can cause mass mortality of host unionid bivalves, and this impact is one of the best-documented negative ecological consequences of dreissenid invasions (reviewed in Karatayev et al. 1997, Lucy et al. 2014). Although fewer studies have been conducted on the impact of D. r. bugensis, it appears that its impact on unionids is far less severe than that of D. polymorpha (Burlakova et al. 2014). In the lower Great Lakes, the replacement of D. polymorpha by D. r. bugensis coincided with significantly lower unionid infestation rates, from an average of >200 dreissenids/unionid in the 1990s (Haag et al. 1993, Gillis and Mackie 1994) to 3 dreissenids/unionid in 2011-2012 (Burlakova et al. 2014). The lower infestation by D. r. bugensis may be due to its lower attachment strength and/or higher species-specific predation by fish compared to D. polymorpha (reviewed in Burlakova et al. 2014).

Impact on fish
The overall impact of Dreissena spp. on fishes depends on the feeding mode of the consumer, the morphology of the invaded water body, the time since the mussel invasion, the coevolutionary history and the Dreissena species, and differs between Europe and North America (Karatayev et al. 2015, Molloy et al. 1997, Strayer et al. 2004, Higgins and Vander Zanden 2010). A total of 77 fish species have been field-documented to feed on adult and/or larval dreissenids on both continents (reviewed in Karatayev et al. in review). In Europe, dreissenids are readily consumed by a wide range of fishes, many of which are evolutionarily adapted to feed on mussels (reviewed in Molloy et al. 1997, Karatayev et al. 2023), and the introduction of dreissenids to new lakes is often associated with increases in fish productivity and commercial catches. In the northern Caspian Sea, the native range of several dreissenid species, approximately 90% of the annual mussel production (13,000 tonnes wet weight) is consumed by fish (Yablonskaya 1985). Roach (Rutilus rutilus) are the most important consumers of several dreissenid species in fresh and brackish waters (Karatayev et al. 1997, Molloy et al. 1997, Karatayev et al. 2023). In addition to roach, dreissenids in Europe are actively consumed by Bream, Silver Bream, European Eel, Russian Sturgeon, Sterlet and Beluga along with 35 other fish species (Karatayev et al. 2023).

Because North American fishes did not co-evolve with dreissenids, early in the Great Lakes invasion, dreissenids were thought to represent a major loss of energy and potential production due to the diversion of food resources from the pelagic to the benthos (Johnson et al., 2005). Dramatic declines in the preferred fish prey, the amphipod Diporeia, caused major losses in the abundance of commercially important fish, including whitefish, alewife, sculpin, bloater and others that are prey for larger piscivores, including salmon and trout (Hoyle et al. 1999, Pothoven et al. 2001, Nalepa et al. 2009a, Karatayev and Burlakova 2022b). As the diet of whitefish shifted from amphipods to Quagga Mussels, the condition, growth and abundance of lake whitefish declined (Pothoven et al. 2001, Hoyle et al. 2008, Nalepa et al. 2009b, Rennie et al. 2009). Over time, however, dreissenids have become an important component of the diets of many commercially important native fishes in invaded North American freshwaters (Karatayev and Burlakova 2022b, Karatayev et al. 2023).

Economic impacts
In most cases, similar to ecosystem impacts, economic impacts were largely identified for D. r. bugensis or for unidentified Dreissena. Dreissena spp. have significant impacts on raw water-dependent infrastructure, including power stations, drinking water treatment plants, industrial facilities, navigation locks and dam structures, and disrupt the operation of pumps, forbays, holding tanks, trashracks and condenser units (Mackie and Claudi 2010; reviewed in Karatayev and Burlakova 2022a). Connelly et al. (2007) estimated the average total economic cost of US power generation and water treatment facilities from 1989 to 2004 at $267 million (range: $161 - $467 million), but these costs do not include costs associated with other impacts on industry and navigation, natural resources (e.g. fisheries), or impacts associated with recreational boating and tourism. The most recent assessment of the global economic costs of dreissenids between 1980 and 2020 was $51.1 billion (2017 US$) (Haubrock et al. 2022), but this estimate suffers from a number of shortcomings, including overlapping costs extracted from different sources (Diagne et al. 2021), the fact that 98% of the data collated were from North America, and the fact that 'cost' categories include several controversial expenditures, such as research, management, detection, surveillance, monitoring, education, communication and information, and risk assessment. Unfortunately, these cost estimates ignore the many benefits of dreissenids, including those to drinking water treatment plants (Wang et al. 2021), water purification, property values, fisheries, etc. (Burlakova et al. 2022, Boltovskoy et al. 2022). such as research, management, detection, surveillance, monitoring, education, communication and information, and risk assessment. Unfortunately, these cost estimates ignore the many benefits of dreissenids, including those to drinking water treatment plants (Wang et al. 2021), water purification, property values, fisheries, etc. (Burlakova et al. 2022, Boltovskoy et al. 2022).

There are no major threats affecting this species. This species itself threatens the biodiversity of the habitats it invades.

Environmental factors controlling D. r. bugensis include substrate and food, eutrophication, pollution, oxygen depletion, competition and predation (reviewed in Karatayev and Burlakova 2022a). In the central basin of Lake Erie at depths >20 m, where bottom hypoxia routinely develops by the end of the growing season, only small numbers of juvenile mussels are found, indicating that hypoxia has eliminated the adult mussels that settled the previous season and limited recruitment and survival of juvenile D. r. bugensis (Karatayev et al. 2018b, 2021b). Therefore, monitoring the distribution of Dreissena can be an effective tool for mapping the extent and frequency of hypoxia in freshwaters. 

The upper temperature limit of D. r. bugensis observed in the field is 30-31°C (Karatayev et al. 1998, 2007, 2014; Wong et al. 2012, Karatayev and Burlakova 2022a). In brackish waters, D. r. bugensis thrives in the least saline areas, from freshwater and up to 3-4‰ (Mills et al. 1996, Orlova et al. 2005, Orlova 2014; Karatayev et al. 2007, 2014).

This species is not known to be used or traded.

Conservation
This species has been given a NatureServe Global Heritage Status Rank of G5 (secure). There are no species-specific conservation measures in place, or needed, for this species, and actions should in fact be taken to limit its spread and population growth, before the negative impacts on many ecosystems become irreversible. Further research is recommended to determine whether this species and D. rostriformis are separate species or two races of the same species.

Research needs
Taxonomic status:
Further research is recommended to determine whether this species and D. rostriformis are separate species or two races of the same species. Although the name D. r. bugensis is the most widely accepted in the literature, further studies are needed to better understand the biological differences between D. r. bugensis and D. rostriformis and to confirm whether D. r. bugensis and the Caspian Sea populations of D. rostriformis are effectively separated by their different salinity tolerances.

Invasion dynamics:
Population explosions early in the invasion have been well documented for D. r. bugensis, but there is a lack of long-term data to predict which environmental factors determine whether the population will fluctuate widely, remain relatively stable for long periods, or decline later in the invasion. More lake-wide long-term studies of dreissenids are needed to understand their dynamics under different environmental conditions, as well as the potential outcomes of long-term coexistence of D. r. bugensis and D. polymorpha.Understand how growth rate and longevity changes over the course of invasion. It is not known whether D. r. bugensis grows faster and lives shorter early in the invasion, when food resources are usually sufficient, than later in the invasion.

Life history:
More data are needed on fecundity, duration of the planktonic stage, time to sexual maturity and sex ratio of D. r. bugensis, as well as the influence of environmental variables such as temperature, depth and food availability. Further work is needed to determine the spawning cues of deep-water D. r. bugensis, as well as their feeding, food spectrum and longevity in the profundal zone of stratified lakes, where their growth is limited by both low temperature and low food concentrations.

Environmental limits:
pH and calcium are among the most important environmental variables limiting the spread of Zebra and probably also Quagga Mussels. While there is a wealth of data on Zebra Mussels, such information on pH and calcium is lacking for Quagga Mussels.

Impact on food webs:
Further studies of the effects of D. r. bugensis on energy flow through food webs in different continents and freshwater habitat types are needed to fully understand and predict its effects on recipient fish communities.

Impact of diseases, predators, and parasites:
Additional research is needed to understand if diseases, predators, and parasites can cause long-term, system-wide declines in D. r. bugensis populations.

Potential ecosystem services an economic benefits:
Currently, the damage caused by D. r. bugensis attracts more attention than its benefits in cost-benefit analyses. Although dreissenids provide considerable benefits, many of which are of great economic importance, their ecosystem and economic benefits are usually ignored, minimised or considered 'non-monetized'. Given the widespread distribution of D. r. bugensis, it is important to quantitatively assess its positive ecological effects and economic benefits as an opportunity to provide additional information to scientists, managers and policy makers.