- Summary of Invasiveness
- Taxonomic Tree
- Distribution Table
- History of Introduction and Spread
- Risk of Introduction
- Habitat List
- Biology and Ecology
- Latitude/Altitude Ranges
- Water Tolerances
- Natural enemies
- Notes on Natural Enemies
- Means of Movement and Dispersal
- Pathway Causes
- Pathway Vectors
- Economic Impact
- Environmental Impact
- Risk and Impact Factors
- Uses List
- Detection and Inspection
- Similarities to Other Species/Conditions
- Prevention and Control
- Gaps in Knowledge/Research Needs
- Principal Source
- Distribution Maps
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PicturesTop of page
IdentityTop of page
Preferred Scientific Name
- Gammarus tigrinus Sexton (1939)
Local Common Names
- Finland: leväkatka
- Germany: bachflohkrebs; gefleckter flussflohkrebs; getigerter
- Netherlands: tijgervlokreeft
- Poland: kielz tygrys
- Russian Federation: bokoplav
- Sweden: tigermärle
Summary of InvasivenessTop of page
Gammarus tigrinus is a small gammarid amphipod naturally occurring in fresh to brackish waters of the East coast of North America. The species was first reported from Europe (UK) in 1931 and later found in the North East Atlantic and in the Baltic Sea, soon being described as one of the most invasive and aggressive alien freshwater species in Europe. It is thought to have been mainly introduced via transportation in ships ballast water and intentional imports as fish food. Its success as an invader is due to a combination of behavioural, physiological and environmental characteristics: it reaches sexual maturity in a short time, produces many generations throughout the season and withstands a wide range of temperatures and salinities. The species has been shown to outcompete and displace native gammarid species and to impose strong predation pressure on other macroinvertebrates, ultimately causing modifications in invaded food web structure and functioning. G. tigrinus has also been responsible for introductions of alien parasites.
Taxonomic TreeTop of page
- Domain: Eukaryota
- Kingdom: Metazoa
- Phylum: Arthropoda
- Subphylum: Crustacea
- Class: Malacostraca
- Subclass: Eumalacostraca
- Order: Amphipoda
- Suborder: Gammaridea
- Family: Gammaridae
- Genus: Gammarus
- Species: Gammarus tigrinus
DescriptionTop of page
G. tigrinus has a typical amphipod body that is segmented throughout, laterally compressed, slender, evenly rounded with a curved or hook-like profile. The body consists of cephalothorax with two pairs of antennas, two eyes and mouthparts; pereon with seven pairs of appendages (pereopods) mainly designed for movement; pleon with another three pairs (pleopods) and urosome carrying three pairs of uropods. Telson is the last segment of the crustacean body. Average body length of a full-grown male is 10.5-12.5 mm, female 8.5-10 mm (Sexton and Cooper, 1939), from the tip of the rostrum to the insertion of telson. Length varies greatly with temperature, salinity and type of food. In Saginaw Bay, Lake Huron, USA, males average 10.5 mm and females average 7.6 mm in length (Grigorovich et al., 2005). Body colour of the living animal is a delicate pale green, yellowish green for males, bluer for females. The remarkable pattern of dark stripes and bars on the light background gave the species the specific name tigrinus. Stripes run across each segment, and two longitudinal bands along the length of the body; the lower is dark, the upper is red. The pattern disappears when specimens are preserved. The cuticle is microscopically spinulose all over, spinules more pronounced in the pleon and in the dorsal region. Mature males G. tigrinus typically have 2-5 groups of posterior marginal setae in the second peduncular (basal) segment of their first antennae, longer and curly second antennae, distinct setae on their pereopods (Sexton and Cooper, 1939; Grigorovich et al., 2005). Uropod 3 is long and setose. Setae are usually less curly during winter months (Jensen, 2010). A detailed description of the segments morphology and their setation can be found in Sexton and Cooper (1939). Overall, morphological differences between male and female in this species are much more noticeable than it is usual for the genus. Females have fewer setae in the antennae and pereopods (Bousfield, 1969; Grigorovich et al. 2005) and are slightly smaller than males.
DistributionTop of page
G. tigrinus is native to shallow lagoons, bays and estuaries along the Atlantic coast of North America, being common from southern Labrador (Canada) to Chesapeake Bay (USA), and sporadic southwards to Florida (Bousfield, 1958; 1963; 1969; 1973). The species has been introduced in different areas of the USA, Europe and Venezuela. Distribution of G. tigrinus tends to show a continuous, rather than fragmented, pattern (Ba et al., 2010). Its southernmost location is in South America, where it has been found in the Gulf of Paria and the Orinoco Delta, Venezuela (Capelo et al., 2004; Martin and Diaz, 2007), and the northernmost one in Finland (harbour of Hamina, Baltic Sea; Pienimäki et al., 2004).
Distribution TableTop of page
The distribution in this summary table is based on all the information available. When several references are cited, they may give conflicting information on the status. Further details may be available for individual references in the Distribution Table Details section which can be selected by going to Generate Report.
|Continent/Country/Region||Distribution||Last Reported||Origin||First Reported||Invasive||Reference||Notes|
|Atlantic, Northeast||Localised||Introduced||1931||Invasive||Kelly et al., 2006a; Bousfield, 1969; Bousfield, 1973; Grigorovich et al., 2005; Kotta et al., 2013||UK, Ireland, Baltic region, Atlantic coast|
|Atlantic, Northwest||Widespread||Native||Kelly et al., 2006a; Bousfield, 1963; Bousfield, 1969; Bousfield, 1973|
|Atlantic, Western Central||Localised||Native||Kelly et al., 2006a; Bousfield, 1969; Bousfield, 1973; Capelo et al., 2004; Grigorovich et al., 2005; Martin and Diaz, 2007||Both native and introduced: native along the USA coast down to Florida; introduced and invasive in the Gulf of Paria and the Orinoco Delta, Venezuela|
|Canada||Widespread||Native||Bousfield, 1958; Bousfield, 1973||From St. Lawrence River, Quebec, southwards|
|USA||Present||Native||Bousfield, 1958; Bousfield, 1969; Bousfield, 1973; Grigorovich et al., 2008; Trebitz et al., 2010||Both native and introduced in the country. Native along the USA coast from Canada to North Carolina. Introduced and invasive in the Great Lakes and the Upper Mississippi River basin|
|-Delaware||Present||Native||Bousfield, 1969||Delaware bay|
|-Florida||Present||Native||Bousfield, 1958; Bousfield, 1963||St John's river and estuary|
|-Maryland||Present||Native||Batie, 1981||Tidal creek of the Rhode River|
|-Michigan||Present||Introduced||Invasive||Grigorovich et al., 2005||Saginaw Bay of Lake Huron in 2002, first record for The Laurentian Great Lakes (although it was present earlier)|
|-New Jersey||Present||Native||Bousfield, 1969||Head of Barnegat Bay|
|-North Carolina||Present||Native||Bousfield, 1969; van Maren, 1978||North River Pond|
|-Virginia||Present||Native||Bousfield, 1969||Chesapeake Bay, Potomac estuary|
|Venezuela||Localised||Introduced||Invasive||Capelo et al., 2004; Martin and Diaz, 2007||Gulf of Paria and the Orinoco Delta|
|Belgium||Widespread||Introduced||1996||Invasive||Messiaen et al., 2010||Eastern part of Flanders|
|Estonia||Present||Introduced||2003||Invasive||Herkul and Kotta, 2007; Herkul et al., 2009; Kotta et al., 2013||Gulf of Riga (Koiguste Bay)|
|Finland||Present||Introduced||2003||Invasive||Pienimäki et al., 2004; Packalén et al., 2008; Sareyka et al., 2011; Kotta et al., 2013||First record in harbours of Hamina and Turku, Gulf of Finland|
|France||Present||Introduced||before 1991||Invasive||Dhur, 1993; Piscart et al., 2005; Piscart et al., 2007; Piscart et al., 2008||First record for France (River Moselle)|
|Germany||Present||Introduced||1957||Invasive||Schmitz, 1960; Bulnheim, 1976; Bulnheim, 1980; Bulnheim, 1985; Rudolph, 1994; Zettler, 1995; Zettler, 2001||First introduction in river Werra|
|Latvia||Present||Introduced||2012||Invasive||Strode et al., 2013||First record for Latvian waters of the Gulf of Riga and in the coastal Lake Liepaja|
|Lithuania||Present||Introduced||2004||Daunys and Zettler, 2006||Curonian Lagoon|
|Netherlands||Present||Introduced||1960||Invasive||Nijssen and Stock, 1966; Pinkster et al., 1977; Pinkster, 1978; Pinkster et al., 1980; Dick and Platvoet, 2000; Josens et al., 2005; Platvoet et al., 2009||First introduction in Ijsselmeer|
|Poland||Present||Introduced||Invasive||Grabowski et al., 2007b; Gruszka, 1995; Gruszka, 1999; Dobrzycka-Krahel et al., 2013||First record for Poland in the Szczecin Lagoon likely in 1988|
|Russian Federation||Present||Introduced||Invasive||Based on regional distribution|
|-Central Russia||Widespread||Introduced||2003||Invasive||Berezina, 2007a; Ezhova et al., 2005; Berezina and Petryashev, 2012; Gusev et al., 2012; Kotta et al., 2013||Gulf of Finland|
|UK||Present||Introduced||1931||Invasive||MacNeil et al., 2003a; Sexton and Cooper, 1939; Costello, 1993; Savage, 2000; Bailey et al., 2006; MacNeil and Dick, 2012||First record for Europe, Worcestershire, Warwickshire and north Gloucestershire|
History of Introduction and SpreadTop of page
Some authors suggest that G. tigrinus was taken to Europe (Ireland) on American naval vessels during World War I and later transported to English estuaries and canals, but this information has never been confirmed (Costello, 1993; Bailey et al., 2006). The first record of G. tigrinus from Europe, in which it was described as a new species, was done by Sexton and Cooper in 1939 (Sexton and Cooper, 1939). However, these authors refer to an earlier observation of the species by Prof. H. Munro in 1931, who found G. tigrinus in abundance in freshwaters contaminated by natural brine seepage near Coventry, in Wyken Slough, and in the brackish waters of River Salwarpe, both in the Droitwich district, Worcestershire. The species was later found in Warwickshire and north Gloucestershire (Sexton and Cooper, 1939). In purely freshwater sites in Northern Ireland, it was found in Lough Neagh, Lough Erne and the lower River Bann, to where it was presumably transported in ballast waters in the first half of the twentieth century (Hynes, 1955; Costello, 1993). Approximately from this period onwards, G. tigrinus started to further colonize European waters, quickly spreading along the Baltic Sea south coast (at rates of around 40 km per year) and covering a total of about 1000 km from 1975 to 1998 (Grigorovich et al., 2005). Despite this quick spread, the species has not yet been observed in Swedish (Främmande Arter, 2006), Danish or Norwegian waters.
Although ballast water has been suggested as a likely introduction vector, the spread of this amphipod in European waters was accelerated by its deliberate release into rivers in Germany and the Netherlands during the 1950-1960s. In 1991, another intentional release of G. tigrinus occurred in River Moselle, in the Germany-Luxembourg border, near Nennig (Massard and Gaby, 1992). Molecular studies have identified multiple introductions from different sources to Europe (Bulnheim, 1985; Pinkster et al., 1992; Kelly et al., 2006b). For instance, the Dutch and German populations originate from Northern Ireland and England, respectively, and these, in turn, may also come from different American source populations.
G. tigrinus from Lough Neagh, Northern Ireland, were introduced into Ijsselmeer, the Netherlands, in 1960 (Nijssen and Stock, 1966; Pinkster et al., 1977). This introduction may have been the result of a semi-intentional release by a scientist who had imported specimens from Northern Ireland for laboratory experiments (Nijssen and Stock, 1966). Between 1964 and 1975, G. tigrinus became the dominant freshwater amphipod in oligohaline waters in the Netherlands (Pinkster, 1978), reaching a more or less stable situation at the end of 1979 (Pinkster et al., 1980). In 1973, the species was found in Haringvliet, the estuary of rivers Meuse and Rhine (Dieleman and Pinkster, 1977). It invaded River Meuse at the latest in 1983 (Berndt, 1984), becoming very abundant at Keizersveer (km 884) in 1991 (Ketelaars, 1993; Ketelaars and Frantzen, 1995) and spreading until Belfeld (km 716). At Grave (km 804), it was the dominant gammarid species from 1992 to 1996, having declined after the arrival of the Ponto-Caspian amphipod Dikerogammarus villosus. It was collected at Borgharen (km 632) in small numbers in 1992, at Petit Lanaye (km 624) in 1995 and at Chokier (km 584) in 1998. In 1984, G. tigrinus was found in the River Rhine and its affluents (Platvoet and Pinkster, 1985) and in the Frisian Isles of Texel, Terschelling and Ameland (Hautus and Pinkster, 1987). It appears that a second invasion, this time coming from German rivers, took place around the 1990s, starting from the eastern part of the country and progressing in a western direction (Pinkster et al., 1992). Later, and mostly due to invasion by the aggressive D. villosus, G. tigrinus was found to be restricted to the Netherlands lowlands and regressing in River Meuse (Dick and Platvoet, 2000; Josens et al., 2005).
The German populations of G. tigrinus were introduced from a brackish lake in England in 1957, as supplement fish feeding to the salt-polluted River Werra, where native gammarids had disappeared (Schmitz, 1960). The species now completely dominates anthropogenically salinized sites of this river (Arle and Wagner, 2013; Szöcs et al., 2014). G. tigrinus spread to the River Weser estuary in 1967, and then east to the oligohaline Kiel Canal, from where it spread to the western Baltic Sea between 1975 and 1979 (Bulnheim, 1980; 1985; Jazdzewski, 1980). In 1975, the species reached the southwestern Baltic Sea, with a record from the westernmost part of Schlei Fjord, Germany (Bulnheim, 1976). After two decades, its rapid spread along the southern Baltic coastline of Germany was noticed, with some authors suggesting it already present in Szczecin Lagoon by 1988 (Wawrzyniak-Wydrowska and Gruszka, 2005; Normant et al., 2007). In 1992, the species was recorded in the Mecklenburg area, more specifically in Peenestrom and Achterwasser in the River Oder estuary (Rudolph, 1994; Zettler, 1995; 2001).
In 1998, G. tigrinus was recorded in the Vistula Lagoon, Poland (Jazdzewski and Konopacka, 2000). From 2001, the gammarid is known to be abundant in northeastern Polish waters, Puck Bay (Gruszka, 2002). The species probably entered Puck Bay from the western part of the Baltic Sea, having moved eastwards along the Polish coastline (Szaniawska et al., 2003). In 2008-2010, G. tigrinus was the most dominant species in nearshore zones of both Vistula Lagoon and Vistula Delta, with native gammarid species becoming absent (Dobrzycka-Krahel et al., 2013). The rapid colonisation of Vistula Lagoon and Vistula Delta might have been facilitated by a salinity increase recorded in the area during the last century (Chubarenko and Tchepikova, 2001; Dobrzycka-Krahel et al., 2013).
In 2004, the species was documented in the Lithuanian part of the Curonian Lagoon (Daunys and Zettler, 2006) and recorded by Ezhova et al. (2005) in waters of Kaliningrad province, Russia. In 2003, it was recorded in Koiguste Bay, northern Gulf of Riga, Estonia (Kotta, 2005; Herkul et al., 2006) where, in 2004, high densities and biomass were registered (Herkul et al., 2009). The Curonian Lagoon in Lithuania is the nearest documented location of G. tigrinus and might be the donor region for the Estonian populations (Herkul and Kotta, 2007). In 2005, the species was found all over the northern Gulf of Riga, throughout the southern coast of Saaremaa Island, and in Rame Bay, western mainland of Estonia. By 2006, G. tigrinus had reached Parnu Bay and the eastern bays of the West Estonian Archipelago Sea. By the end of 2008, the species had been found in a total of 350 benthos samples from the Estonian coastal sea (Herkul et al., 2009). In Latvia, it was recorded during field surveys in 2011-2012, from different sites of the Gulf of Riga and in the coastal Lake Liepaja (Strode et al., 2013).
In 2003, the gammarid was found in Finnish waters (Pienimäki et al., 2004), in harbours of Turku, northern Baltic, and in Hamina, Gulf of Finland, the latter being the northernmost record for the species. In 2005, G. tigrinus was recorded in the eastern-most part of the Baltic Sea, near a new oil terminal in the River Neva estuary, Russia, where, in 2006, its density had more than quadrupled (Berezina, 2007a). Spreading eastwards along the estuary, at a speed of about 100 km per year (Berezina, 2007a), the species reached the northern and southern parts of the estuary by 2006-2009. In 2011, it was recorded from Luga and Koporye Bays in the south, and from the northern part of the Russian Gulf of Finland (Berezina and Petryashev, 2012). By 2013, G. tigrinus had become established in shallow water habitats of the whole coastal zone of the Gulf of Finland (Kotta et al., 2013).
In Belgium, the first record of G. tigrinus in the wild is from 1996, in a pond near Antwerp (Vercauteren et al., 1996; Kerckhof et al., 2007). However, in 1991, the species was already reported in canals (Messiaen et al., 2010), which indicates that it might have arrived earlier via the Scheldt-Rhine connection and other canals. It is possible that the introduction into Belgium occurred through natural expansion via River Meuse, or by accident while replacing fish stocks from Dutch into Belgian waters. By 2010, the species had invaded part of the Kleine Nete and Postelvaartje.
In northeastern France, G. tigrinus was first recorded in 1991 (Dhur, 1993) in River Moselle, an oligohaline tributary of River Rhine. Since then, the species has rapidly dispersed via canals that connect to other river catchments, such as rivers Saône and Rhône (in 1995; Fruget and Dessaix, 2003) and the upper Loire River (in 2003; Bollache et al., 2004). In 2002, the species was sampled from Meurthe River (Piscart et al., 2005). A clear range expansion was observed when the species reached River Vilaine and the Nantes-Brest canal (Piscart et al., 2007; 2008). G. tigrinus was first reported along the southern coast of Brittany in 2005 (Piscart et al., 2007). By September 2008, its range had expanded by more than 60 km to the North and West of that observed in 2005 (Piscart et al., 2008). The species is currently spreading at a slow rate along the Atlantic coast of France (Piscart et al., 2005; 2007; 2008). Whereas the eastern part of France seems to have been invaded by a freshwater tolerant strain of G. tigrinus (Piscart et al., 2008), the distribution and dispersal of the species in Brittany suggests a recent introduction to the region, probably via Saint Nazaire harbour (River Loire estuary) and from a brackish water origin (e.g. ballast water transfer). G. tigrinus of brackish origin are expected to slowly spread into freshwaters, with an increased dispersal rate once populations have adapted to lower salinity levels (Kelly et al., 2006b).
G. tigrinus was introduced in the American Great Lakes in 2002, possibly with ballast water transported from Europe (Grigorovich et al., 2005; Kelly et al., 2006b). It was initially found in Saginaw Bay, Lake Huron, but archived material indicates the species was present in Lake Superior and Lake Erie already in 2001. Furthermore, according to Trebitz et al. (2010), the species was already present in St. Louis River/Duluth, Lake Superior, in 1985. Subsequent collections reveal the species has broadly colonized shallow-water habitats around the perimeter of the Great Lakes, being now, in many wetland habitats, the second most abundant amphipod after G. pseudolimnaeus (Grigorovich et al., 2005). In 2004-2006, zoobenthic surveys in the Upper Mississippi River basin (Missouri, Upper Mississippi and Ohio rivers) revealed new invasions by euryhaline amphipods (Grigorovich et al., 2008). G. tigrinus and Echinogammarus ischnus were discovered in the lower reaches of the Ohio and Upper Mississippi rivers in 2004, with an increase in range and abundance from 2004 to 2006. The presence of breeding adults suggests the species is permanently established in this area.
Recent records by Capelo et al. (2004) and Martin and Diaz (2007) suggest the species has started to invade the Western Central Atlantic, where it was found in the Gulf of Paria and the Orinoco Delta, Venezuela.
Risk of IntroductionTop of page
Many authors highlight the risk of G. tigrinus invading new waterbodies as very significant. In Europe, for example, there is a high risk of further expansions to various lakes of Eastern Europe via inland canal-river systems (Berezina, 2007a,b), to lentic waters of Brittany and most French large waterways (Piscart et al., 2007; 2008), and along the Baltic coast, especially under ongoing environmental changes related to increased eutrophication and warming climate (Lenz et al., 2011; Sareyka et al., 2011; Kotta et al., 2013).
Ba et al. (2010) used ecological niche modelling to analyse the invasive potential of G. tigrinus. The authors predict the global occurrence of the invasive amphipod to expand mainly through shipping. Ports with salinity between 29.54 and 35.66 in Argentina, Australia, Brazil, Canada, China, Cote d'Ivoire, Japan, New Zealand, North Korea and Far East Russia are at high risk.
HabitatTop of page
In its native range, G. tigrinus is essentially benthic, common in the upper portion of estuaries in both fresh and brackish waters (Bousfield, 1973; van Maren, 1978). It lives in debris, among algae (Enteromorpha [Ulva]) and hydroids (Cordylophora), on stakes and pilings, hard substratum or sand. It is a low intertidal (often on rocky shores in beach seeps and stream outflows), but essentially subtidal species. Populations are found in a wide range of inland and coastal ecosystems in temperate regions, indicating low habitat selectivity in its native range.
In the American invaded range, the species mostly occurs in very shallow areas on silty sand, on Cladophora or in macrophyte beds (Grigorovich et al., 2005). In the European invaded range, G. tigrinus seems to have a narrower niche, thought to be a consequence of its restricted genetic pool (Kelly et al., 2006b; Kotta et al., 2013; 2014). The abundance and biomass of G. tigrinus vary as a function of wave exposure, water salinity and transparency, being higher at less exposed, less saline and more turbid sites (Kotta et al., 2014). Accordingly, at the type location (UK), the species was found in sheltered places at the edge of the river among leaves, roots and debris around Glyceria aquatica [Catabrosa aquatica] plants, or from undersides of stones (Sexton and Cooper, 1939). In Northern Ireland, inland populations of G. tigrinus occur in deep, mid-Lough areas and, to a limited extent, in sheltered, vegetated shores (Dick, 1996). Similarly, in the northern Baltic, the species inhabits charophyte communities from sheltered to moderately exposed coasts, where salinity is low, eutrophication moderate-high and visibility moderate to poor (Herkul et al., 2009; Kotta et al., 2014). The species prefers sheltered sandy and silty sediments associated with charophytes, phanerogams (Myriophyllum spicatum and Potamogeton pectinatus [Stuckenia pectinata]) and the green alga Cladophora glomerata. It is suggested that the dense charophyte communities provide food and, due to high structural complexity, serve as refuge from predators (Kinzler and Maier, 2003; Orav-Kotta and Kotta, 2004; Kotta et al., 2013; 2014). In Poland, the species inhabits the littoral nearshore zone bordered by rush (Phragmites sp., Typha sp., Scirpus sp., sometimes Pilayella and Ulva), among submerged plants and in sandy or sandy-muddy bottoms with decaying plant remains (Dobrzycka-Krahel et al., 2013). In the Lithuanian part of the Curonian Lagoon, it was found in all types of habitats sampled (reeds, mixed and soft bottoms), except in enclosed depositional environments with mixed substrates and mud present (Daunys and Zettler, 2006). In the invaded range, habitat differentiation may take place to avoid intraguild predation and cannibalism (Dick, 1996; van Riel et al., 2007; 2009). A field survey of a Dutch lake invaded by both G. tigrinus and Dikerogammarus villosus showed D. villosus to be dominant on the rocky boulder substrate of the shoreline and G. tigrinus to dominate the crushed shell/sand matrix immediately adjacent to this (MacNeil et al., 2008). In Lough Neagh, Northern Ireland, a similar scenario exists between G. duebeni celticus and G. tigrinus, with G. d. celticus dominating the boulder substrate in the nearshore areas of the lake and G. tigrinus the sandy substrates of the mid-Lough (MacNeil and Prenter, 2000; MacNeil et al., 2003b).
Habitat ListTop of page
|Estuaries||Principal habitat||Harmful (pest or invasive)|
|Lagoons||Principal habitat||Harmful (pest or invasive)|
|Irrigation channels||Secondary/tolerated habitat||Harmful (pest or invasive)|
|Lakes||Principal habitat||Harmful (pest or invasive)|
|Ponds||Secondary/tolerated habitat||Harmful (pest or invasive)|
|Reservoirs||Secondary/tolerated habitat||Harmful (pest or invasive)|
|Rivers / streams||Principal habitat||Harmful (pest or invasive)|
|Rivers / streams||Principal habitat||Natural|
|Rivers / streams||Principal habitat||Productive/non-natural|
|Coastal areas||Principal habitat||Harmful (pest or invasive)|
|Coastal areas||Principal habitat||Natural|
|Intertidal zone||Principal habitat||Harmful (pest or invasive)|
|Intertidal zone||Principal habitat||Natural|
Biology and EcologyTop of page
Kelly et al. (2006a) suggest that deep levels of population structure among American G. tigrinus populations developed consistently with ancient glaciation events. Analysis of mitochondrial cytochrome c oxidase I and nuclear internal transcribed spacer 1 phylogenies support this deep genetic structure and the existence of major northern and southern phylogroups. The occurrence of divergent population structure of G. tigrinus supports limited dispersal of the species and highlights the importance of historical factors as determinants of estuarine speciation and diversification of this gammarid in its native range. This strong native phylogeographical structure allowed Kelly et al. (2006b) to identify three allopatrically evolved clades in invading populations of G. tigrinus and to determine multiple introductions in Europe. The study shows that the most divergent clades occur in the British Isles and in mainland Europe, and were sourced from St. Lawrence and Chesapeake/Delaware Bay estuaries. A third clade was found in the Great Lakes, USA, and sourced to the Hudson River estuary. Patterns of habitat colonisation by different native clades suggest differences in salinity tolerance among lineages. G. tigrinus did not occur in freshwater at putative source sites, but colonized both fresh and brackish waters in the British Isles, Western Europe and Eastern Europe. Populations consisting of admixtures of two of the invading clades were mainly found in recently invaded fresh and brackish water sites in Eastern Europe, and were characterized by a higher genetic diversity than the source populations.
Reproductive biology of the species has been studied in detail by Chambers (1977). G. tigrinus undergoes direct development, with no independent larval stage. Females carry their embryos in a brood chamber between the pereopods. When released, juveniles reach maturity after several moults, without any metamorphosis. Sexual maturity is reached at a small size, with females only slightly over 4 mm having been observed bearing eggs. Maturity is usually achieved in a short time period, but this is greatly influenced by temperature. For example, Hynes (1955) stated that it took 13 weeks for the species to reach sexual maturity at 16°C and 6 weeks at 25°C. According to Pinkster (1975; 1978), maturity was reached in 40 days at 10°C, 32 days at 15°C and 27 days at 20°C. Savage (1982) suggested that G. tigrinus reaches sexual maturity after 8.7 weeks at 15°C, and after 6 weeks at 25°C.
Reproduction can be triggered at any time of the year, if temperature increases by 15-16°C (Ginn et al., 1976). Males grasp females in a characteristic pre-copula pose, and hold on until spawning takes place. Overwintering animals may copulate in winter, and egg-bearing females have been found in early spring at a water temperature of 2°C. Females can produce 3-16 generations during each reproductive season, with the exception of the coldest months, when water temperatures are close to 0°C. There are usually three spawning peaks per year, in July, August and October. The number of eggs in a brood depends on the size of the female, but mean brood size is about 20 eggs. The maximum number of eggs recorded in a brood is 95 (Grabowski et al., 2007a). In Saginaw Bay, Great Lakes, USA, reproductive females carry, on average, 32 embryos and populations contain many more juveniles and females than males, probably because males have shorter life spans (Grigorovich et al., 2005).
In the invaded range, G. tigrinus shows higher fecundity, more generations per year, the ability to mature earlier and reproduce throughout the year. This effective reproduction strategy may help outcompete native gammarids, such as G. duebeni and G. zaddachi that have slower reproduction tempo (Pinkster et al., 1977; Savage, 1982; Sareyka et al., 2011).
Life span of G. tigrinus is about 2 months, depending on temperature. Overwintering animals may live for several months.
Population Size and Structure
Population densities of G. tigrinus vary from some individuals to extremely high numbers, depending on season, location and the invasion stage. In inland European waters, the highest density was described for reed beds of Lake Tjeukemeer, the Netherlands, where the species reached more than 20,000 individuals per m2 (Chambers, 1977). In the southeast Baltic, from 2008 to 2010, mean abundance of G. tigrinus increased from 33 to 234 ind/m2 in Vistula Lagoon, and from 7 to 95 ind/m2 in Vistula Delta. G. tigrinus have the highest biomass in Vistula Lagoon (0.48 g/m2), reaching nearly 100% frequency among other gammarid species (Dobrzycka-Krahel et al., 2013). In the northern Baltic, Estonian coastal sea, G. tigrinus reached maximum biomass in Koiguste Bay in 2004, its second year of invasion (Herkul et al., 2009). By the end of 2008, species abundance varied between 0.118 and 5150 ind/m2 and biomass from <0.001 to 12.22 g/m2, with average values of 418 ind/m2 and 0.654 g/m2, respectively. In the Gulf of Riga, abundance ranges between 5 and 100 ind/m2 (Strode et al., 2013).
Gammarid amphipods are generalist foragers, very flexible in utilising diverse food sources (Poje et al., 1988; Kotta et al., 2014). G. tigrinus filter feeds on suspended organic matter and can also directly consume zooplankton, plants, algae and detritus (Hunte and Myers, 1984; MacNeil and Prenter, 2000; Grigorovich et al., 2005; Normant et al., 2007). The species is an important aquatic shredder that prefers to ingest and digest leaves colonized by aquatic hyphomycete fungi (Barlocher and Porter, 1986; Sridhar and Barlocher, 1993; Sridhar et al., 2011).
G. tigrinus can also be described as a facultative carnivore (Savage, 2000). In Watch Lane Flash, UK, and in water bodies in the Netherlands and Germany, G. tigrinus has been noted to be a voracious predator (e.g. Schmitz, 1960; Fries and Tesch, 1965; Chambers, 1977; 1987; Savage, 1982; Pinkster et al., 1992). Predation by G. tigrinus upon other amphipods has been observed and contributes to the decline of native gammarids. Laboratory microcosm experiments show that G. tigrinus cannibalism, particularly by large size individuals upon smaller ones, may be common (MacNeil et al., 2008). G. tigrinus preys on relatively small Crangonyx pseudogracilis in Northern Ireland, and could similarly prey on it in the Great Lakes (Dick, 1996; Grigorovich et al., 2005).
In the Baltic area, G. tigrinus has often been found associated with Dreissena polymorpha beds (Radziejewska et al., 2009). Other invertebrate species frequently co-occurring with G. tigrinus are chironomid larvae, the cockle Cerastoderma glaucum and the gastropod Theodoxus fluviatilis (Orav-Kotta and Kotta, 2004; Kinzler and Maier, 2003; Kotta et al., 2013; 2014).
Lenz et al. (2011) collected G. tigrinus individuals from shore habitats dominated by the brown alga Fucus vesiculosus, and Kotta et al. (2014) associate the species with Chara aspera macrophyte communities.
G. tigrinus survives temperatures close to 0°C and upper temperature tolerance is 32-34°C (Kipp, 2007). Survival of G. tigrinus at higher temperatures is more successful in ion-rich, polluted waters, indicating a wider physiological tolerance and higher competitive ability of the species in these habitats (Wijnhoven et al., 2003). In laboratory experiments, G. tigrinus survived after 10 h both at 30 and 32°C, while 50% of G. zaddachi individuals, a native gammarid species, died after an average of 5.6 h and 1 h, respectively. Furthermore, while 50% of all G. tigrinus individuals survived at 35°C for an average time span of 5.6 h, all G. zaddachi individuals died immediately (Lenz et al., 2011). In the Baltic area (Vistula Delta and Lagoon), the species inhabits waters where temperature varies between 11.4 and 24°C (Dobrzycka-Krahel et al., 2013).
G. tigrinus is highly euryhaline (Devin and Beisel, 2007; Jensen, 2010). In its native range, it lives in brackish waters of salinities from 4 to 20 ppt (Kelly et al., 2006b) and tolerates a range of 0 to 25 ppt (Bousfield, 1973). In Europe, the lower critical salinity level for survival is 0.2 ppt, and the upper level 7.1 ppt in the British Isles and 16-18 ppt in the rest of Europe, with physiological variations between geographically distinct populations (Ruoff, 1968; Savage, 1982). In the Baltic area, the species has been found in salinities of 0.2 to 4.9 ppt (Dobrzycka-Krahel et al., 2013).
Oxygen saturation levels below 30% are lethal to all experimental populations of amphipods. In the Back and Savannah River estuaries, Georgia, USA, G. tigrinus is especially susceptible to low dissolved oxygen levels, exhibiting mortality within 3 h of exposure to levels between 12 and 18% (Winn and Knott, 1992). Still, laboratory experiments show that G. tigrinus is more resistant to hypoxia than G. zaddachi (Sareyka et al., 2011).
Tolerance to poor water quality enables G. tigrinus to establish in areas characterized by deteriorated environmental conditions (Savage, 1996; MacNeil et al., 2001; 2007). For example, in the southern Baltic Sea, the species has successfully colonized severely eutrophicated estuaries (Szaniawska et al., 2003; Jazdzewski et al., 2004; 2005; Wawrzyniak-Wydrowska and Gruszka, 2005) and, in the Rhine estuary, deteriorated conditions promote an increase in G. tigrinus populations (Platvoet and Pinkster, 1995).
In its introduced area, the species often lives in waters with elevated ionic content due to pollution (Streit and Kuhn, 1994; Koop and Grieshaber, 2000; Wijnhoven et al., 2003; Grabowski et al., 2007a,b). G. tigrinus is able to regulate extracellular potassium ions in river waters with high potassium concentrations, and to tolerate stress levels of sodium and chloride ions (>380 mmol dm−3), making it more successful in salt-polluted rivers than other amphipods (Koop and Grieshaber, 2000). Additionally, Streit and Kuhn (1994) found that G. tigrinus is more tolerant to organophosphorus insecticides than G. pulex and G. fossarum, giving the former a survival advantage in the Rhine River. However, G. tigrinus is highly sensitive to heavy metals (Strode and Balode, 2013), cadmium (Boets et al., 2012) and excess natural organic matter (Timofeyev et al., 2006).
Latitude/Altitude RangesTop of page
|Latitude North (°N)||Latitude South (°S)||Altitude Lower (m)||Altitude Upper (m)|
Water TolerancesTop of page
|Parameter||Minimum Value||Maximum Value||Typical Value||Status||Life Stage||Notes|
|Salinity (part per thousand)||4||20||Optimum||Tolerated: 0-25 ppt|
|Water temperature (ºC temperature)||15||20||Optimum||Tolerated: 0-34°C|
Natural enemiesTop of page
Notes on Natural EnemiesTop of page
Fishes, including perch Perca fluviatilis, stickleback Gasterosteus aculeatus and eel Anguilla anguilla, birds and dragonfly nymphs are known to feed on G. tigrinus. Fish can cause heavy mortality on G. tigrinus populations but, in turbid environments, G. tigrinus may escape heavy predation because benthivorous fish locate their prey visually (e.g. Kotta et al., 2010; 2013). Dikerogammarus villosus, another invasive gammarid, predates all size classes of G. tigrinus, with the smallest size classes suffering the highest predation (MacNeil et al., 2008).
In invaded waters, prevalence of parasites and epibionts (such as rotifer Embata parasitica and the protozoan Epistylis) seems to be higher in the native G. duebeni celticus than in invading amphipods, including G. tigrinus, which may increase the species invasive success (Dunn and Dick, 1998). Fungus Saprolegnia sp. affects G. tigrinus (Dieleman and Pinkster, 1977; van der Velde et al., 2000) and the parasite Maritrema subdolum (Trematoda) has been reported to occur in G. tigrinus in Poland (Rolbiecki and Normant, 2005).
Means of Movement and DispersalTop of page
A wide geographical spread in Finnish, Estonian and Russian waters, within a short time period, probably indicates several introductions and successful establishments, rather than gradual species migration (Sareyka et al., 2011). Accordingly, Grigorovich et al. (2008) believe the natural dispersal capabilities of G. tigrinus are insufficient to account for their range expansion in the large rivers of the Upper Mississippi basin, USA.
The main human-mediated means contributing to the expansion of G. tigrinus are unintentional and consist of transport in ship ballast water (Sylvester and MacIsaac, 2010; Sareyka et al., 2011), construction of new waterways, webs of small waters and harbours at river mouths (van der Velde et al., 2000; Wooten et al., 2006), and escapes from aquaculture (Nijssen and Stock, 1966). Some other means of amphipod dispersal, such as hull fouling (Sylvester and MacIsaac, 2010) and algal (Myers, 1993) or animal (Swanson, 1984) rafting, seem to pose a low risk of G. tigrinus introduction and spread.
Pathway CausesTop of page
|Aquaculture||Escapes from aquaculture facilities||Yes||Nijssen and Stock, 1966|
|Habitat restoration and improvement||In Germany, as supplement fish feeding in polluted habitats where native gammarids disappeared||Yes||Yes||Schmitz, 1960|
|Intentional release||Semi-intentional release in the Netherlands||Yes||Yes||Nijssen and Stock, 1966|
|Interbasin transfers||Mainly in ship ballast waters||Yes||Yes||Grigorovich et al., 2005; Grigorovich et al., 2008; Kotta et al., 2013; Kotta et al., 2014|
|Interconnected waterways||Spreading via waterways, including canals, in France, Germany and USA||Yes||Yes||Piscart et al., 2007; Grigorovich et al., 2008; Piscart et al., 2008|
Pathway VectorsTop of page
|Floating vegetation and debris||Potential to disperse on floating seaweed||Yes||Myers, 1993|
|Host and vector organisms||Potential to disperse on aquatic animals||Yes||Swanson, 1984|
|Ship ballast water and sediment||Main mean of introduction to new areas||Yes||Yes||Grigorovich et al., 2005; Sylvester and MacIsaac, 2010; Sareyka et al., 2011|
|Ship hull fouling||Possible||Yes||Sylvester and MacIsaac, 2010|
|Water||Dispersal by water currents or through connected waterways||Yes||Yes||van der Velde et al., 2000; Wooten et al., 2006|
Economic ImpactTop of page
When in high densities, G. tigrinus can cause damage to fishing gear and trapped fish (Pinkster et al., 1977).
Environmental ImpactTop of page
Impact on Biodiversity
In areas where G. tigrinus and other non-native gammarids dominate, the original amphipod fauna has often been drastically changed (e.g. Pinkster et al., 1992; Jazdzewski et al., 2004; Zettler, 2008). In the Baltic Sea, for example, native gammarid species of the Vistula Lagoon G. zaddachi and G. duebeni, which were common in 1952-1956 and in 1970, have been gradually replaced by G. tigrinus and other non-indigenous gammarids, such as Pontogammarus robustoides, Dikerogammarus haemobaphes and Obesogammarus crassus (Zmudzinski, 1957; Jazdzewski, 1975; Jazdzewski et al., 2004; Grabowski et al., 2006). In the Vistula Delta, the native species G. duebeni, G. zaddachi, G. salinus and G. oceanicus, common in 1967-1971, had mostly been replaced in 1998-2000 by the non-indigenous species D. haemobaphes, P. robustoides and G. tigrinus (Jazdzewski, 1975; Jazdzewski et al., 2004). In 2008-2010, G. tigrinus was the most dominant species in nearshore zones of both Vistula Lagoon and Vistula Delta (Dobrzycka-Krahel et al., 2013), while native gammarid species went totally extinct. Similarily, following the invasion of G. tigrinus, G. salinus almost disappeared from Koiguste Bay, Estonia (Herkul et al., 2009) and G. duebeni from Watch Lane Flash, UK (Savage, 1982). In Northern Ireland, G. tigrinus may be more disadvantaged from intraguild predation by the native G. duebeni celticus than vice versa, but the native species appears more disadvantaged with respect to drift, parasitism and the interaction of the two (MacNeil et al., 2003b,c).
One of the reasons for this native gammarid retreat to occur is the fact that G. tigrinus proved to be competitively superior to the majority of native gammarids in Europe (Pinkster et al., 1992; Grabowski et al., 2006; Orav-Kotta et al., 2009; Kotta et al., 2006; 2010; Sareyka et al., 2011). Due to its ecological plasticity, aggressiveness and fast reproduction, G. tigrinus competes successfully for food and space with native gammarid species, forcing them to leave their preferred habitats and increasing exposure to fish predation (Orav-Kotta et al., 2009; Kotta et al., 2010; 2011). For example, in the Gulf of Finland, the omnivorous G. tigrinus is able to outcompete and replace native herbivorous Gammarus species (Packalén et al., 2008). In River Rhine, interspecific competitive stress caused by G. tigrinus, combined with physiological stress caused by the ion-enriched Rhine water, may have restricted the distribution of G. pulex to ion-poor ditches and brooks discharging into this river (van Riel et al., 2009). It is important to note that, for some regions, drastic changes in salinity, and not the introduction of G. tigrinus, are the most important factor causing native gammarid decline (Piscart et al., 2005; Boets et al., 2011; Kotta et al., 2014).
Being an active carnivore, G. tigrinus shows significant predatory impact on other macroinvertebrates. In Germany, the impact was noted by Fries and Tesch (1965), in the Netherlands Roos (1979) reported the species feeding on polyps of Cordylophora caspia and in Ireland G. tigrinus coexists with the native opossum shrimp Mysis relicta through mutual predation (Bailey et al., 2006). Females and recently moulted M. relicta are particularly vulnerable to predation by G. tigrinus, which has forced the species to change its microhabitat use, exposing itself more to fish predation.
G. tigrinus invasions have also been accompanied by the simultaneous introduction of alien parasites (Leppakoski, 2002). G. tigrinus has been shown to be an intermediate host to the acanthocephalan Paratenuisentis ambiguus, which definitive host is the American eel Anguilla rostrata (Samuel and Bullock, 1981; Taraschewski et al., 1987; Morozinska-Gogol, 2008; 2009). This acanthocephalan seems to have been introduced with G. tigrinus into Europe, where its final European host is A. anguilla (Thielen, 2005; Taraschewski, 2006). During the 1990s, both exotics had spread into River Rhine, with G. tigrinus dominating the amphipod fauna and P. ambiguus dominating the intestinal helminth community of local eels (Sures and Streit, 2001). However, after the inauguration of the Rhine-Main-Danube Canal in 1992 and subsequent invasion by the competitively superior D. villosus, G. tigrinus was eliminated and P. ambiguus stopped being the dominant intestinal helminth of the European eel, not being recorded since 2001 (Dick and Platvoet, 2000; Thielen, 2005). The parasite has recently been found in eels in Polish coastal waters (Morozinska-Gogol, 2008) and may also occur in other localities within G. tigrinus introduced range.
Although G. tigrinus actively cleans water from detritus and dead animals, it is not able to compensate for the activity of native species, which replacement has led to a decrease of at least 66% in the rate of leaf litter recycling (van der Velde et al., 2000, Piscart et al., 2011). Furthermore, despite being a food source for commercial fish, when abundant, the species becomes included in the food web, causing changes and shifts in fish diet, which results in overall changes of the littoral food web (Kelleher et al., 1998; 2000; Grabowski et al., 2006). Recent experimental evidence indicates that G. tigrinus has a strong potential to modify benthic community structure and functioning in invaded coastal zones (Orav-Kotta et al., 2009; Kotta et al. 2010; 2011; 2013; Sareyka et al., 2011).
Risk and Impact FactorsTop of page Invasiveness
- Invasive in its native range
- Proved invasive outside its native range
- Has a broad native range
- Abundant in its native range
- Highly adaptable to different environments
- Is a habitat generalist
- Tolerates, or benefits from, cultivation, browsing pressure, mutilation, fire etc
- Pioneering in disturbed areas
- Capable of securing and ingesting a wide range of food
- Fast growing
- Has high reproductive potential
- Has high genetic variability
- Altered trophic level
- Damaged ecosystem services
- Ecosystem change/ habitat alteration
- Modification of natural benthic communities
- Modification of nutrient regime
- Negatively impacts aquaculture/fisheries
- Reduced native biodiversity
- Threat to/ loss of native species
- Competition - monopolizing resources
- Pest and disease transmission
- Interaction with other invasive species
- Rapid growth
- Highly likely to be transported internationally accidentally
- Difficult to identify/detect as a commodity contaminant
- Difficult to identify/detect in the field
- Difficult/costly to control
UsesTop of page
G. tigrinus is considered an important fish food, especially under conditions of depletion of native food sources. In the Lithuanian coastal zone, the amphipod was found to contribute to up to 40% of the diet of some commercially important fish species (Bubinas, 1994). Several species, including perch Perca fluviatilis, stickleback Gasterosteus aculeatus and eel Anguilla anguilla, feed on G. tigrinus (Daunys and Zettler, 2006). Nellis et al. (2007) considered G. tigrinus a potential food source for sturgeon in St. Lawrence estuarine transition zone harbours, Canada, where the only known concentrations of age-0 and early juveniles of St. Lawrence Atlantic sturgeon and lake sturgeon populations exist.
Uses ListTop of page
Animal feed, fodder, forage
- Live feed
Detection and InspectionTop of page
The species is difficult to detect because specialist expertise and a microscope are needed to identify it. Amphipod species can be identified using taxonomic keys by Bousfield (1958; 1989), Holsinger (1976) and Grigorovich (1989).
Similarities to Other Species/ConditionsTop of page
Many species of Gammarus and related genera (Chaetogammarus, Dikerogammarus, Echinogammarus, Obesogammarus, Pontogammarus) look very similar. In all gammarids, antenna I has article 2 more than half as long as article 1, which is not thickened and body pigmented and, in the Gammarus genus, article 3 of antenna I is not longer than article 1 (Bousfield, 1958; 1973). Classification of amphipods from genus Gammarus strongly depends on life stage and gender-specific characters, such as: 1) shape of the interantennal lobe of the head; 2) setosity of the peduncular and flagellar segments of antennae I and II; 3) shape and armature of pereopods V; and 4) armature of the epimeral plates (Sexton and Cooper, 1939; Bousfield 1958; 1969; Holsinger, 1976).
Several native and introduced gammarid species occur in the same general area as G. tigrinus. For how to distinguish G. tigrinus from other Baltic species, such as G. duebeni and G. zaddachi, see Jensen (2010); for other European species see Dobson (2012) and NNSS (2014). For American gammarids, distinctions between local species have been described by Bousfield (1969) and Grigorovich et al. (2005).
Prevention and ControlTop of page
Particular emphasis should be placed on preventing human-mediated dispersal of G. tigrinus, especially in high invasion risk ports (Ba et al., 2010). Water of such ports should be regularly monitored for exotic aquatic organisms that may survive there temporarily or permanently. An European surveillance network and an appropriate adaptation of monitoring strategies are recommended, with the aim to assess risks, impacts and the invasiveness of alien species, as well as the efficiency of eradication and control measures (Nehring and Klingenstein, 2005). Regulatory international measures against invasions in European Seas are discussed in Gollasch (2006).
Development of simple identification keys and leaflets are recommended, aiming to raise public awareness about G. tigrinus.
Eradication does not seem to be possible.
Gaps in Knowledge/Research NeedsTop of page
Further investigation on the distribution, reproductive success and ecological significance of G. tigrinus are needed, in order to quantitatively estimate its role in the structure and functioning of benthic communities in European waters. MacNeil and Dick (2012) recommend future studies to examine changes in interaction strengths among amphipods and mysids under rising temperature and eutrophication, which currently affect European freshwaters.
Petruck and Stoffler (2011) pointed out that G. tigrinus proposed salinity preference does not correlate with the one detected in River Lippe, Germany, which may require further investigation.
ReferencesTop of page
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Principal SourceTop of page
Draft datasheet under review
ContributorsTop of page
23/09/14 Original text by:
Ekaterina Shalaeva, Consultant, UK
Distribution MapsTop of page
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