An overview of the distribution and ecology of the alien cyanobacteria species Raphidiopsis raciborskii, Sphaerospermopsis aphanizomenoides and Chrysosporum bergii in Europe

Raphidiopsis raciborskii is a globally distributed filamentous diazotrophic cyanobacterium and the most studied invader among cyanobacteria. It was originally described as Anabaena raciborskii by Wołoszyńska from Java (Indonesia, 1899-1900). However, polygenetic analyses have suggested central African lakes as the original centre of dispersion (Vico et al. 2020). The species was considered a typical tropical species (Wołoszyńska et al. 1912), but it also occurs in lakes and water reservoirs in the temperate zone, including Europe, where conditions may be different from their previously known habitats (Fastner et al. 2003, Briand et al. 2004, Mankiewicz-Boczek et al. 2012).

According to Padisák (1997), this species was first found in Europe in 1938, in Lake Kastoria in Greece (Fig. 1; Table S1). In the 1960s and 1970s, the species was observed in Belarus (Mikheeva 1967), Ukraine (Hamar 1977, Horecká & Komárek 1979, Présing et al. 1996), Hungary (Padisák 1997 and references therein), Austria (Claus 1961, Padisák 1997 and references therein), Slovakia (Horecká & Komárek 1979) and the Czech Republic (Horecká & Komárek 1979). A sudden increase of new reports in Europe indicated that this was a new invader (Padisák 1997). In 1973, the species was also reported from northern Europe, namely from a highly modified ecosystem – Lake Pątnowskie, a strongly heated lake used for cooling a power plant in Poland (Burchardt 1977), and later in Lake Jieznas in Lithuania in 1988 (Kavaliauskienė 1996). Later still, in the 1990s, R. raciborskii was found in the water body near St. Petersburg (Russia) (Balashova et al. 1999), 7 water bodies in Germany (Krienitz & Hegewald 1996, Fastner et al. 2003, Mischke 2003, Nixdorf et al. 2003, Stüken et al. 2006, Botanic Garden & Botanical Museum Berlin 2021), a gravel-pit lake, Janíčkov dvor, in Slovakia (Maršálek et al. 2000), a fishpond, Mézeshegyi-tó, in Hungary (Borics et al. 2000), several water bodies in Romania (Cărăuş 2012), 3 water bodies in France (Coûté, Leitao & Martin 1997, Briand et al. 2002), Lake Volvi and the River Strymon in Greece (Padisák, 1997 and references therein), and 10 water bodies in Portugal (Saker 2004). There was a considerable increase of new reports in the first two decades of the 21^st century: in Lake Zazari in Greece (Vardaka et al. 2005), the River Seine and water bodies from Viry-Châtillon, Chanteraînes, Courneuve in France (Druart & Briand 2002; Gugger et al. 2005), Lakes Albano, Trasimeno, Cedrino, and Biviere di Gela in Italy (Manti et al. 2005, Barone et al. 2010), Lake Albufera in Spain (Romo et al. 2008 and references therein), 41 lakes in the Berlin-Brandenburg region in Germany (Stüken et al. 2006, Haande et al. 2008, Botanic Garden & Botanical Museum Berlin 2018), at least 24 new localities in Poland (Stefaniak & Kokociński 2005, Kokociński et al. 2009, Budzyńska & Gołdyn 2017), Lake Sakadaš in Croatia (Mihaljević & Stević 2011), Slatina pond and Reservoir Aleksandrovac in Serbia (Cvijan & Fužinato 2012, Simić et al. 2014), 11 waterbodies in the Netherlands (Knoben & Wal 2020), Reservoir Kasperivtsi and the Rivers Seret, Dnierp and Don (Ukraine) (Tsarenko et al. 2006, Kaštovský et al. 2010 and references therein; Rzymski et al. 2018) and Lake Nero in Russia (Babanazarova et al. 2015). 23 localities were reported in the Czech Republic (Kaštovský et al. 2010) and 22 localities in Bulgaria in 2000-2015 (Stoyneva-Gärtner et al. 2017). According to Komárek (2013), R. raciborskii has spread northward to Finland.

Figure 1

R. raciborskii has gradually become the dominant species in Lake Balaton, Hungary, causing blooms several times since 1982 with a maximum chlorophyll-α ranging from 70 to 160 μg l^-1, and later in 1992 with a biomass up to 870 mg l^-1 in a shallow hypertrophic fishpond in Mézeshegyi-tó (Présing et al. 1996, Borics et al. 2000) (Fig. 1, Table S1). During the years 1993 and 1994, R. raciborskii was reported in a shallow, wind mixed urban Lake Alte Donau, in Vienna, Austria, where it contributed up to 90% of the total phytoplankton biovolume (Dokulil & Mayer 1996). During the years 1992, 1995 and 1999, this species reached up to 97% of the total phytoplankton biomass in the Danubian wetlands in Bulgaria (Stoyneva 2003). In 1999, the species accounted for more than 99% of the total phytoplankton density in the Francs–Pêcheur pond in France (Briand et al. 2002). In the years from 1994 to 2004, R. raciborskii reached more than 10% of the phytoplankton biomass in Lake Kastoria in Greece (Katsiapi et al. 2013). In addition, the seasonal abundance of R. raciborskii reached up to 3 x 10⁶ cells ml^-1 in Reservoirs Odivelas, Caia, Maranhão (Portugal) in 1999 (Saker et al. 2004). Due to its adaptation to summer temperatures, prolonged water retention, and wide range of nitrogen levels, the species began to dominate in oligohaline Lake Albufera (Spain) from 1998 to 2006 (Romo et al. 2008). In 2004, R. raciborskii reached up to 50% of the total phytoplankton biomass in Italian Lake Albano (Messineo et al. 2010) and in the years from 2005 to 2007 a change in phytoplankton composition was observed in the southern Italian Lake Biviere di Gela due to a high abundance of R. raciborskii, which dominated the phytoplankton along with Chrysosporum ovalisporum and Pseudoanabaena limnetica (Barone et al. 2010). According to Barone et al. (2010), the transformation was due to increased air temperature and evapotranspiration, which reduced the water inflow and disrupted the littoral zone of the lake. R. raciborskii accounted for 24% of the total phytoplankton biomass in Lake Scharmutzelsee in Germany in 1999 (Nixford et al. 2003). In the years from 2005 to 2007, R. raciborskii was found in Reservoir Rusałka in Poland, where it accounted for up to 79.4% of the total phytoplankton biomass (Budzyńska & Gołdyn 2017). In 2008, it was observed in shallow, turbid, eutrophic Polish lakes (Lakes Żabiniec, Szydłowskie, Niepruszewskie, Bnińskie), where the cyanobacterial biomass accounted for up to 13.9% of the total phytoplankton biomass. Kokociński and Soininen (2012) found that the biomass of R. raciborskii was positively correlated with the total nitrogen, total phosphorus, and conductivity, while negatively correlated with the temperature, ammonium, and orthophosphate. In 2014, this species was found in 24 out of 101 investigated lakes in Poland, where it accounted for 0.09%–24.60% of the total phytoplankton biomass (Kokociński et al. 2017b). The first bloom of R. raciborskii in Serbia was observed in the eutrophic lowland river Ponjavica in 2008. The species accounted for more than 85% of the total phytoplankton biomass (Karadžić et al. 2013). In 2018, it accounted for 5–25% of the total phytoplankton density in Reservoir Poroy and Lake Uzungeren in Bulgaria (Stefanova et al. 2020).

According to the Köppen-Geiger climate classification system and a historical overview of the distribution of R. raciborskii in Europe, the species first colonised ecosystems in the Mediterranean climate (Fig. 1, Table S1) (Padisák 1997, Peel et al. 2007). Already in the 1970s, this species had spread to the northern part of Europe (Russia) (Balashova et al. 1999). Since then, R. raciborskii has significantly expanded its range, establishing itself in various climatic zones such as cold semi-arid, humid subtropics, temperate oceanic, and even in the tundra and subarctic zones. Most reports were in the humid continental climate zone in Central Europe. The northermost locations noted so far are in northwestern Russia (Lake Nero, 57°09′09.9″N 39°26′24.1″E) and Finland (unspecified location) (Komárek 2013, Babanazarova et al. 2015). To date, R. raciborskii has been most commonly reported from shallow eutrophic lakes, but also occurs in deep lakes, mesotrophic lakes, mountain lakes, artificial reservoirs, fish ponds, marshes, wetlands, rivers, and river deltas in at least 176 water bodies in twenty-one countries in Europe. The species is widely distributed in Portugal, the Netherlands, Poland, Germany, and Bulgaria, although it may be related to intensive research in these countries. R. raciborskii is not only a widespread species, but also a bloom forming species. It has produced blooms in 10 countries where it accounted for up to 99% of the total phytoplankton biomass.

Sphaerospermopsis aphanizomenoides (Forti) Zapomelová, Jezberová, Hrouzek, Hisem, Reháková & Komárková 2010

Syn.: Aphanizomenon aphanizomenoides (Forti) Hortobágyi & Komárek 1979, Anabaena aphanizomenoides Forti 1911, Aphanizomenon sphaericum Kisselev 1955

Sphaerospermopsis aphanizomenoides was originally described from Lake Anatolia in Turkey in 1911 (Geitler 1932), but has also been reported from tropical and subtropical regions in Asia (Malaysia, India) and South America (Brazil) (Desikachary 1959; Prowse 1972; see references in Horecká & Komárek 1979; Bittencourt-Oliveira et al. 2011) (Fig. 2; Table S2). In the second half of the 20th century, S. aphanizomenoides was found in warmer areas of Europe and from 1950 till the first decade of the XXI century has been noted in Hungary (Hortobagyti 1955, Borics et al. 2000), 32 sites in the Czech Republic (Horecká & Komárek 1979, Zapomělová et al. 2012), the Danube River in Romania (Cărăuş 2012), and Lake Stará Morava in Slovakia in 1997 (Hindák 2000). At the beginning of the 21st century, S. aphanizomenoides was observed in Lake Vela in Portugal (de Figueiredo et al. 2010), 3 water bodies in France (Brient et al. 2009, Ledreux et al. 2010), Lake Doirani in Greece (Vardaka et al. 2005), a pond in Casas de Millán, and Reservoir Montijo and River Guadiana in Spain (Moreno et al. 2005, Wörmer et al. 2011). In Serbia, this species was first detected in the Ponjavica River in 2002 (Karadžić et al. 2013), and six years later it accounted for 24% of the total phytoplankton biomass (Jovanović et al. 2015). S. aphanizomenoides was reported in at least 15 water bodies in Poland and formed blooms in three of them (Reservoir Rusałka, Lakes Swarzędzkie and Uścimowskie) (Stefaniak & Kokociński 2005, Kokociński & Soininen 2012, Budzyńska & Gołdyn 2017, Budzyńska et al. 2019). It also has been found in 10 water bodies in Germany (Stüken et al. 2006), 6 water bodies in the Netherlands (Knoben & Wal 2020), in the United Kingdom (John et al. 2002), the hypereutrophic Cabras lagoon in Italy (Pulina et al. 2011), Reservoir Kasperivtsi and River Seret in Ukraine (Rzymski et al. 2018), and Lakes Jieznas, Gauštvinis and Širvys in Lithuania (Karosienė et al. 2020). It is considered invasive in Spain, Poland, Germany, and the Czech Republic (Stefaniak & Kokociński 2005, Stüken et al. 2006, Kaštovský et al. 2010).

Figure 2

S. aphanizomenoides was originally described from the Mediterranean climate in the Asian part of Turkey (Geitler 1932). While its native habitat is considered to be near the European border, it has recently expanded its distribution in Europe significantly. It has become established in various climatic zones such as the Mediterranean, humid subtropical, and temperate oceanic climates. It has never been found in cold semi-arid, tundra or subarctic climates. Similar to R. raciborskii, most reports have been found in the humid continental climate zone in Central Europe. The northernmost location noted so far is in Lithuania (Lake Gauštvinis, 55°38′54.0 ″N 23°11′35.6 ″E) (Karosienė et al. 2020). In total, the species has been recorded in 15 countries in at least 83 water bodies. It formed blooms in 4 water bodies where it accounted for up to 62% of the total phytoplankton biomass.

This cyanobacterium may be even more common than recorded due to its morphological similarity to the trichomes of Aphanizomenon gracile when lacking akinetes (Zapomělová et al. 2012).

Chrysosporum bergii (Ostenfeld) Zapomelová, Skácelová, Pumann, Kopp & Janecek 2012

Syn.: Anabaena bergii Ostenfeld 1908

Chrysosporum bergii and the morphologically similar Chrysosporum minus were originally distributed in the habitats of the Aral Sea and the Ponto-Caspian region, which includes the Caspian Sea, Lake Issyk-Kul, the Black Sea, lakes near the Ural Mountains, and relict lakes in Central Asia (Elenkin 1938; Gollerbakh et al. 1953; Proshkina-Lavrenko & Makarova 1968). Several strains are known from Africa and Australia (Stüken et al. 2009; Koreivienė & Kasperovičienė 2011). Both species were found in the brackish Danube delta, which is the oldest record of their occurrence in Europe (Vladimirova & Danilova 1968). Later, they have also been reported from Serbia (Cvijan & Krizmanić 2009; Simić et al. 2014), and the Netherlands (Veen et al. 2015), but the specificities of species are poorly described. Only Chrysosporum minus has been described in Austria (Hindák & Deisinger 1989; Hindák 1992). Chrysosporum bergii has also been reported from the River Dyje (the Czech Republic) (Heteša et al. 1997), the gravel pit lake near Trávnik (Slovakia) (Hindák 2000), one water body in Belarus (Mikheeva 1999), and more recently in the Marathonas reservoir (Greece) (Katsiapi et al. 2011). This species occurred in 13 out of 142 German water bodies studied in 2004, although Anabaena bergii var. minor Kiselev and Anabaena bergii f. minor were considered synonyms to Anabaena bergii Ostenfeld by these authors (Stüken et al. 2006). There are numerous reports describing the occurrence of C. bergii in Poland (Kokociński et al. 2013, Budzyńska & Gołdyn 2017, Budzyńska et al. 2019, Kokociński & Soininen 2019). In 2012 C. bergii occurred in 7 out of 19 randomly selected lakes in Poland. Therefore, it was considered as a common species in this region, although it has never formed blooms (Kokociński & Soininen 2019). The latest record of this species is from Lake Vaya in Bulgaria, where it was found in 2018 (Stefanova et al. 2020).

Anabaena minderi was first described in the monomictic Lake Greifen and 4 ponds near Zurich in Switzerland (Huber-Pestalozzi 1938, Couté & Preisig 1978) and under the name Anabaena bergii var. limnetica in the reservoirs of the Seine and Marne rivers in France (Couté & Preisig 1978) (Fig. 3; Table S3). This taxa has been found in the Trávnik gravel pit lake from the Komárno district (Slovakia) (Hindák 2000), Lake Piasecno (Poland) (Bucka & Wilk-Woźniak 2005), Lakes Šmartinski, Pernijško and Grajševsko (Slovenia) (Remec-Rekar et al. 2008) and in the Netherlands (Veen et al. 2015). The northernmost locality of Anabaena minderi is the most recent record in Europe (Gineitiškės Lake in Lithuania in 2008) (Koreivienė & Kasperovičienė 2011).

Figure 3

Chrysosporum bergii and Chrysosporum minus were originally described from the brackish habitats of the Caspian and Aral Seas and tend to invade northern European areas that are less brackish than their original habitats (Zapomelová 2012). In contrast, the morphologically similar Anabaena minderi was first described in alpine lakes with low salinity in Switzerland (Huber-Pestalozzi 1938, Komárek 2013). Therefore, probably A. minderi should be considered native to European waters. However, its taxonomy is unclear and should be revised. All three species have expanded their distribution into the humid continental climate, as well as temperate oceanic and humid subtropical zones. A. minderi occurs also in the tundra climate. The northernmost locations noted so far are in Germany for C. bergii (53°10′51.9″N 13°14′33.8″E) (Stüken et al. 2006), in the Netherlands for C. minus (unspecified location) (Veen et al. 2015), and in Lithuania for A. minderi (Lake Gineitiškės, 54°44′12.4″N 25°11′06.9″E) (Koreivienė & Kasperovičienė 2011). In total, the species have been recorded in 15 countries in at least 29, 5 and 14 water bodies for C. bergii, C. minus and A. minderi, respectively. They have never formed blooms in Europe.

Temperature is widely considered an important factor for the growth of R. raciborskii and its spread. However, both field and laboratory studies show the species can tolerate a wide range of temperatures. A study of 28 mesotrophic and hypereutrophic lakes from tropical, subtropical, and temperate climates reports its presence in lakes with temperatures as low as 11°C (Bonilla et al. 2012). The laboratory experimentation of Briand et al (2004) showed that both tropical and temperate strains were able to grow in a temperature range of 20 to 35°C, with the maximum growth rates at around 30°C. Culture and microcosmos experiments of isolates from water bodies in Germany showed very similar results, with the optimal temperature for this species being around 28°C and the maximum temperature of growth at 35°C. It revealed that R. raciborskii had significantly higher growth rates than native species at high temperatures (≥ 20°C), while having lower growth rates at low temperatures (≤ 15°C) (Mehnert et al. 2010). The tolerance of a wide range of temperatures (11 to 35°C) may explain the ability of R. raciborskii to establish itself in temperate areas. This species is expected to spread further, and increase its proliferation and competitiveness under global warming scenarios.

The laboratory experiments showed that S. aphanizomenoides tolerates a wide range of temperatures (from 10 to 40°C) (Sabour et al. 2009a, Mehnert et al. 2010). However, the maximum growth rates presented in two studies varied: 35°C (Sabour et al. 2009a) and 29°C (Mehnert et al. 2010), probably due to differences in the culturing methods. Even though the species is reported to be able to grow in low temperatures in laboratory conditions, it is only found in the warmest period of the year in the temperate climate, contrary to R. raciborskii, reported also form colder months. Similarly to R. raciborskii, S. aphanizomenoides is a much better competitor at higher temperatures compared to native species (Mehnert et al. 2010). Savadova et al. (2018) showed that under laboratory conditions S. aphanizomenoides was most favoured at temperatures of 20 to 30°C, but the optimal temperatures varied between the strains. A bloom formation of S. aphanizomenoides occurred in the small, shallow and eutrophic Ponjavica River in Serbia during warm months of the year. A field study showed the correlation between the species’ biomass and high water temperatures (Jovanović et al. 2015). Another field study conducted in a Polish reservoir, reported a bloom of the species in an exceptionally hot summer. According to the authors, the succession of this species was highly dependent on the water temperature (Budzyńska & Gołdyn 2017). Therefore, warming is among the factors promoting this species spread and its establishment in temperate zones.

In contrast to R. raciborskii and S. aphanizomenoides, C. bergii was a better competitor among nostocalean species at moderate water temperatures (19–20°C) and the maximum temperature for growth of the species was 26.5°C under laboratory conditions (Mehnert et al. 2010). Two field studies conducted in Gineitiškės Lake in Lithuania for A. bergii var. limnetica and in 19 Polish lakes for C. bergii revealed that these species developed in the warmest months, when the water temperature ranged from 20.4 to 23°C and 21.8 to 23.9°C, respectively (Koreivienė & Kasperovičienė 2011, Kokociński et al. 2019). New data from the culture experiment showed that C. bergii had its maximum growth rates at 30°C and it decreased at lower temperatures (18–20°C) (Savadova et al. 2018). The variability observed in the results concerning the temperatures associated with the maximum growth rates of C. bergii indicates the presence of distinct temperature preferences. This variation could be attributed to the diversity among strains of C. bergii or the lack of consensus in the taxonomy of morphologically similar species. Therefore, further studies are needed to clarify the optimal temperature for each species discussed in this paper and to determine whether these species differ in their ecology.

The overview of different culture and field experiments revealed that the optimal intensity for R. raciborskii growth varied between 50 and 150 μmol photons m^-2 s^-1, but optimal illumination highly differs among the different strains (Burford et al. 2016 and references therein). This species could grow under low-light conditions (10 μmol photons m^-2 s^-1) in laboratory conditions, therefore, it is considered a shade tolerant species (Pierangelini et al. 2014). Some strains of R. raciborskii show photoinhibition under light intensities above 100 m^-2 s^-1 (Briand et al. 2004), while another laboratory study reports high growth rates in photon fluxes of 348 μmol photons m^-2 s^-1 (Carneiro et al. 2013). The diversity of light preferences may suggest the existence of high- and low-light adapted ecotypes. Furthermore, studies have shown that R. raciborskii is shade-tolerant due to its ability to increase the concentration of phycobiliproteins that absorb the broader light spectrum (Pagni et al. 2020).

The light intensity required for the growth of S. aphanizomenoides isolates was at moderate light intensities (100–110 μmol photons m^-2 s^-1) (Mehnert et al. 2010). Another laboratory study showed that the photoinhibition of this species is related to water temperature. Growth of S. aphanizomenoides at a temperature of 15°C was saturated at a light intensity of up to 130 μmol photons m^-2 s^-1. In contrast, at a temperature of 35°C, the optimal light intensity was reached at up to 442 m^-2 s^-1 (Sabour et al. 2009a). Therefore, this species may have a low light preference at low water temperatures. In addition, both S. aphanizomenoides, C. bergii and A. bergii var. limnetica were mostly found in water bodies with low transparency (Stüken et al. 2006, Tezanos & Litchman 2009, Budzyńska et al. 2019). These results suggest that all the species mentioned here tolerate low light conditions. This is very beneficial as the blooms themselves decrease the water transparency. Besides this, the cold period and low light intensities are among the most stressful factors for cyanobacterial species inhabiting temperate surface waters (Burford et al. 2016).

The ability to exploit different light conditions is one of the factors contributing to the species succession. Another advantage for the species adaptation is their ability to settle at different depths due to gas vacuoles and to regulate their position in the water column to the most favourable position (Padisák 1997, Antunes et al. 2015).

Orthophosphate limitation in nature has led to the development of various metabolic strategies that allow cyanobacteria to cope with it. R. raciborskii tolerates low environmental phosphorus concentrations due to its high affinity to dissolved inorganic phosphorus (DIP) and its storage capacity (Isvánovics et al. 2000). Thus, the species can store dissolved organic phosphorus (DOP) in the cytoplasm in the form of polyphosphates and utilizes excess internal phosphate under conditions of low DIP in the environment (Burford & Davis 2011). A lake experiment showed that the addition of DIP increased the dominance of this species (Posselt et al. 2009). Therefore, R. raciborskii prefers environments with high phosphorus concentrations, but is not dependent on it and can tolerate a wide range of DIP.

S. aphanizomenoides is a highly nutrient-demanding species. The biomass of this species correlated positively with the presence of a high phosphate concentration both in water bodies (Budzyńska et al. 2019) and under laboratory conditions (Savadova-Ratkus et al. 2021). Budzyńska et al. (2019) suggested phosphorus as the primary driving factor of this species growth, and eutrophication as a factor that stimulates the expansion of S. aphanizomenoides towards higher latitudes. The species is able to uptake organic phosphorus under phosphorus limited conditions using alkaline phosphatase (Cirés & Ballot 2016). In contrast, the biomass of C. bergii correlated negatively with DIP (Kokociński & Soininen 2019). This species occurred more frequently in shallow lakes with low dissolved phosphorus amounts (Kokociński & Soininen 2019).

R. raciborskii dominates in water bodies with different nitrogen concentrations and mostly prefers ammonium as the nitrogen source (Burford & O’donohue 2006, Antunes et al. 2015, Engström-Öst et al. 2015). Energetically, dissolved inorganic nitrogen (DIN) is a more efficient source for R. raciborskii growth than atmospheric nitrogen (N₂). For this reason, the uptake rates of DIN are higher than the nitrogen fixation rates. Atmospheric nitrogen fixation gives an advantage that depends on the nitrogen content in the cell (Sprőber et al. 2003).

S. aphanizomenoides is another N₂ fixing cyanobacterium. (Sabour et al. 2009b) found a positive correlation between the species biomass and nitrogen concentration, but nitrogen depletion did not limit the growth of the species. Budzyńska and Gołdyn (2017) showed that the biomass of this species is positively related to high ammonium-nitrogen concentrations. Savadova-Ratkus et al. (2021) observed that inorganic nitrogen significantly affects the growth rate of S. aphanizomenoides, but no effect on diazotrophic C. bergii was found. Hindák and Hindáková (2001) reported that the occurrence of C. bergii and A. minderi in Slovak lakes may be related to eutrophication.

The results indicate that R. raciborskii is a species tolerant to a wide range of nutrient concentrations. On the other hand, phosphorus concentration is a limiting factor for the growth of S. apahanizomenoides. In contrast, nutrients are not among the factors limiting the growth of C. bergii, which may even prefer low inorganic phosphorus concentrations.

The concentration of carbon dioxide, pH, and salinity are among the variables important for cyanobacteria proliferation. CO₂ concentration in the water doesn’t significantly affect the environmental performance of R. raciborskii (Antunes et al. 2015). It can dominate at a pH between 6.9 and 10 (Saker 2004, Antunes et al. 2015), thus, it can tolerate high pH values due to its ability to use carbon sources other than CO₂, which decreases in alkaline environments (Burford et al. 2016). R. raciborskii prefers low salinity environments, especially oligohaline and mesohaline conditions (0.5-8 ppm) (Antunes et al. 2015, Engström-Öst et al. 2015). High salinity (> 30 ppm) suppressed growth of this species (Moisander et al. 2012), but at high nutrient concentrations the species was able to dominate even in saline waters (Engström-Öst et al. 2015). C. bergii tolerates brackish environments (Zapomělová 2012). The existing literature does not currently include any studies that analyse the potential influence of salinity or pH on S. aphanizomenoides.

The survival of alien species in winter of the colder part of the temperate zone may be explained by the tolerance to low temperatures and the formation of akinetes (resting cells) which are found in most European heterocytous cyanobacterial populations (Padisák 1997, Saker 2004). In contrast, the formation of akinetes in tropical strains is rare. They can persist in the vegetative form throughout the year (Saker & Griffiths 2000). Akinetes formation also promotes survival on dispersal routes (Saker 2004). In Europe, akinetes germination is promoted by warm conditions and therefore high temperatures may enable stronger incoming blooms as a result of higher phosphorus accumulation (Budzyńska & Gołdyn 2017).

A great metabolic plasticity or existing ecotypes may explain the success of R. raciborskii under various conditions (Briand et al. 2004, Piccini et al. 2011, Bonilla et al. 2012, Kokociński et al. 2017a, Pagni et al. 2020). Metabolic plasticity is a rapid adaptive mechanism under changing environmental conditions since a single genotype can produce different phenotypes by altering its gene expression. Although plasticity is a useful mechanism for becoming established in a new environment, it is energetically disadvantageous. Metabolic plasticity demands energy resources to maintain genetic information, produce the phenotype, acquire information about the environment, and keep the phenotype flexible in relation to the changing environment (Pigliucci 2001). Metabolically plastic species have multiple genes for specific factors; on the contrary, ecotypes of a single species have specific genes for a particular factor and can survive in one type of the environment only due to the presence of these genes. Ecotypes are functionally distinct groups that are genetically and phenotypically stable (Burford et al. 2016; Baxter et al. 2020). Ecotypes differ in their morphology, physiology, toxins, and genetics, and are adapted to local climatic conditions. It is not entirely clear whether the global distribution of R. raciborskii is due to high metabolic plasticity, which allows rapid adaptation to different environments, or to the existence of distinct ecotypes that allow to thrive in specific environments (Pagni et al. 2020). Pagni et al. (2020) suggest that distinct ecotypes are the most likely hypothesis, as many studies have confirmed limitations on cyanobacterial growth and development due to genetic differences between strains. Comparison of genetic information of strains isolated from several different ecosystems in a gradient of environmental variables is urgently needed to explain this phenomenon.

One of the factors influencing the establishment of cyanobacteria in the new aquatic communities is their ability to produce allelochemicals. These allelochemicals influence the phytoplankton’s competition, succession, and bloom formation (Fistarol et al. 2004). Alien cyanobacteria can produce potentially harmful compounds, although this ability is strain-specific (Kokociński & Soininen 2012). R. raciborskii is known to produce cylindrospermopsin (CYN) and saxitoxin (STX) and its analogues (paralytic shellfish poison, PSP). CYN was first detected in Australia, when it caused an outbreak of hepatoenteritis among indigenous people (Ohtani et al. 1992). Later, CYN production was also observed in New Zealand (Wood & Stirling 2003). STX and PSP, on the other hand, are produced by R. raciborskii in South America (Vico et al. 2020).

Contrary to the reports from Australia, Oceania and South America, none of the R. raciborskii strains isolated so far from Europe, Africa and North America, has been found to produce known cyanotoxins (Haande et al. 2008, Yilmaz & Phlips 2011, Antunes et al. 2015, Vico et al. 2020). However, some of the R. raciborskii strains isolated from Portugal, Hungary, Poland and Germany were demonstrated as toxic to various invertebrate or vertebrate models, even though they did not contain any of the known toxins (Fastner et al. 2003, Saker 2004, Antal et al. 2011, Ács et al. 2013, Poniedziałek et al. 2015). The presence of a specific cyrJ gene was detected in a CYN-positive natural bloom sample from a pond in the Czech Republic dominated by both Raphidiopsis raciborskii and R. mediteranea. However, toxicity was not confirmed in cultures, and the presence of the gene could also be attributed to some native CYN-producing species (Blahova et al. 2021). In addition, microcystin synthetase genes were present in a strain of R. raciborskii isolated from Lake Karla (Greece), but toxin production was not confirmed by chemical analyses (Panou et al. 2018).

Synthesis of cyanotoxins depends on environmental variables. The production of CYN by R. raciborskii was negatively correlated with temperature. Maximum growth of this species was reached at high temperatures (35°C), while it produced CYN at lower temperatures (20°C), so blooms may have low toxicity at high temperatures (Saker & Griffiths 2000). Therefore, if toxic strains become established in Europe, the colder climate could affect its toxicity. In addition, a positive relationship between light intensity and CYN production was reported by Dyble et al (2006).

Sabour et al. (2005) noted that the strain of S. aphanizomenoides isolated from the shallow brackish Lake Oued Mellah in Morocco produced microcystins, but toxicity studies of this species have not been genetically confirmed. S. aphanizomenoides has occurred in toxic blooms in Europe, but toxin production by the species has also never been confirmed (Wörmer et al. 2011, Cirés & Ballot 2016, Karosienė et al. 2020). Two nostocalean strains morphologically similar to Sphaerospermopsis were positive for cyrB and cyrC genes, and ESI-LC-MS/MS confirmed CYN production, but phylogenetic analysis of 16S rRNA indicated that they likely belong to a different genus (Cordeiro et al. 2021).

Chrysosporum bergii was confirmed as a potential CYN producer in Australia (Schembri et al. 2001), but CYN-producing strains of C. bergii were reclassified as C. ovalisporum in Europe (Stüken et al. 2009). Toxic strains of this species have never been found in Europe.