Aquatic Invasions (2014) Volume 9, Issue 3: 267–288
doi: http://dx.doi.org/10.3391/ai.2014.9.3.04
© 2014 The Author(s). Journal compilation © 2014 REABIC
Open Access
Proceedings of the 18th International Conference on Aquatic Invasive Species (April 21–25, 2013, Niagara Falls, Canada)
Review
The profile of a ‘perfect’ invader – the case of killer shrimp, Dikerogammarus villosus
Tomasz Rewicz 1 * , Michal Grabowski 1 , Calum MacNeil 2 and Karolina Bącela-Spychalska 1
1
Department of Invertebrate Zoology and Hydrobiology, University of Lodz, 12/16 Banacha, 90-237 Lodz, Poland
Department of Environment, Food and Agriculture, The Isle of Man Government, Thie Slieau Whallian, Foxdale Road, St. Johns IM4 3AS,
Isle of Man, The British Isles
2
E-mail: tomek.rewicz@gmail.com (TR), michalg@biol.uni.lodz.pl (MG), Calum.MacNeil@gov.im (CML), karolina@biol.uni.lodz.pl (KBS)
*Corresponding author
Received: 26 May 2014 / Accepted: 7 July 2014 / Published online: 2 August 2014
Handling editor: Vadim Panov
Abstract
The ‘killer shrimp’, Dikerogammarus villosus, has been recognised as one of the 100 worst alien species in Europe, in terms of negative
impacts on the biodiversity and functioning of invaded ecosystems. During the last twenty years, this Ponto-Caspian amphipod crustacean
has rapidly spread throughout Europe’s freshwaters and its invasion and continued range expansion represents a major conservation
management problem. Although a great deal of research has focused on this almost ‘perfect’ invader as its damaging impacts, realised and
potential, have become evident, we now present the first comprehensive review of D. villosus taxonomy, morphology, distribution,
community impacts, parasites, life history, physiological tolerance and finally, possible eradication methods. We show the direct and indirect
ecosystem impacts of this invader can be profound, as it is a top predator, capable of engaging in a diverse array of other feeding modes. It
can quickly dominate resident macroinvertebrate communities in terms of numbers and biomass, with subsequent large-scale reductions in
local biodiversity and potentially altering energy cycling, such as leaf litter processing. This damaging European invader has the potential to
become a key invader on a global scale as it may be capable of reaching North American freshwaters, such as the Great Lakes. One positive
aspect of this invader’s spread and impact is increased interest in alien species research generally, from decision-makers, stakeholders and
the general public. This has resulted in greater financial support to study invasion mechanisms, preventative measures to stop invasion spread
and ways to minimise damaging impacts. Our review provides a specific example, that studies identifying management strategies that
mitigate against a potential invader’s spread should be undertaken at the earliest possible opportunity in order to minimise potentially
irreversible ecosystem damage and biodiversity loss.
Key words: biological invasions, non-indigenous species, Amphipoda, Ponto-Caspian, risk assessments, aquatic invasive species (AIS)
Introduction
Alien species represent a major threat to
conservation management on both a continental
and global scale (Leppäkoski et al. 2002; Chandra
and Gerhadt 2008; Richardson and Ricciardi 2013).
Invasion by alien species is increasingly recognised
as one of the major threats to biodiversity in freshwater ecosystems (Sala et al. 2000; Holdich and
Pöckl 2007; SCBD 2010; Lambertini et al. 2011).
The ‘killer shrimp’, Dikerogammarus villosus
(Sowinsky, 1894), is a euryoecious amphipod
crustacean of Ponto-Caspian origin, regarded as
one of the worst one hundred invasive species in
Europe (DAISIE 2009). It is a highly voracious,
physiologically tolerant and adaptable species,
threatening freshwater biodiversity and ecosystem
functioning on various levels (Bollache et al. 2008;
MacNeil et al. 2010; Piscart et al. 2010). Within
two decades, it has succeeded in colonising most
of the major European inland waterways replacing
many resident amphipod ‘shrimp’ species, including
previously successful invaders (Bij de Vaate et
al. 2002; Bollache et al. 2004; Grabowski et al.
2007c; Bącela et al. 2008). In 2011 the species
was detected outside mainland Europe for the
first time, namely in Great Britain, in an English
reservoir called Grafham Water (MacNeil et al.
2012). Subsequently, several more populations
were detected in quick succession in other parts
of England and Wales (Environmental Agency
2012; MacNeil et al. 2012). If D. villosus spread
follows the pattern of many other aquatic invaders,
its range could expand beyond Europe to
267
T. Rewicz et al.
Figure 1. The number of
publications dealing with various
aspects of Dikerogammarus
villosus invasion, published up to
2013 and registered in the
SCOPUS database.
eventually reach North American freshwaters, such
as the Great Lakes, as has previously happened
with the zebra (Dreissena polymorpha (Pallas,
1771)) and quagga mussels (Dreissena rostriformis
bugensis (Andrusov, 1897)), as well as another
amphipod, Echinogammarus ischnus (Stebbing,
1899) (Ricciardi and MacIsaac 2000). Taking into
account the ever increasing amount of research
that the scientific community has focussed on the
killer shrimp, we conducted the first comprehensive
review of the literature, including searching the
Scopus database with the keywords Dikerogammarus villosus and ‘killer shrimp’, as well as
sourcing unpublished reports and local Russian
literature. We thus aim to both summarise and
critically evaluate all the major published studies
dealing with this species, its invasion history,
ecology, interaction with local communities, its
invasion potential and issues of control and
eradication (Figure 1).
Taxonomic remarks and potential reasons for
the success and range expansion of PontoCaspian invaders in European watercourses
According to Bousfield (1977), genera such as
Dikerogammarus, Pontogammarus, and Obesogammarus, are grouped in the family Pontogammaridae and all include species which are
invasive to numerous parts of Europe. The
taxonomy of the Dikerogammarus genus remains
a source of both confusion and contention. For
instance, some 12 species are ascribed to this
genus in various papers (summarised by Grabowski
et al. 2011), but as a consequence of weak species
268
definition and loss of type materials, further
taxonomic revisions are required that will probably
reduce this number. However, more recent studies
have also revealed the presence of cryptic species
within the genus (Grabowski and Jażdżewska
unpublished data), adding further to the confusion.
Fortunately, for this current review, among the
Dikerogammarus species, Dikerogammarus villosus
is the most completely defined in morphological
terms. An established and comprehensive literature
allows for its unambiguous identification among
other congeneric species alien in Central and
Western Europe, such as Dikerogammarus haemobaphes (Eichwald, 1841) or Dikerogammarus
bispinosus Martynov, 1925 (Eggers and Martens
2001; Konopacka and Jażdżewski 2002; MordukhaiBoltovskoi et al. 1969; Özbek and Özkan 2011).
Contrary to the singular opinion of Pjatakova and
Tarasov (1996) that D. villosus should be
synonymised with D. haemobaphes, other studies
have shown the clear differentiation of D. villosus
from D. haemobaphes and also from D. bispinosus,
both on morphological and molecular bases
(Carauşu 1943; Carauşu et al. 1955; Müller et al.
2002; Wattier et al. 2006) (Figure 2).
Dikerogammarus villosus belongs to the ‘PontoCaspian faunistic complex’, which includes predominantly euryoecious animal species, originally
endemic to the coastal areas of the Caspian Sea,
Azov Sea, Black Sea, their brackish lagoons
(limans), and associated lower reaches of rivers
which drain to these seas (Mordukhai-Boltovskoi
1964; Stock 1974; Jażdżewski 1980; Barnard and
Barnard 1983). The Ponto-Caspian basin resulted
from the transformation of the Neogene
Dikerogammarus villosus - profile of the invader
Distribution in native and colonized range
Native range of Dikerogammarus villosus
Figure 2. Live specimen of Dikerogammarus villosus (picture
by Michal Grabowski).
epicontinental Sea of Parathetys (Dumont 2000).
Tectonic movements followed by sequential
regressions/transgressions then transformed the
sea into a number of brackish and freshwater
lakes. Since the beginning of the Pleistocene (ca.
2.5 Mya), rapid climatic and geological changes
resulted in reformation of the Ponto-Caspian basin.
This in conjunction with glaciation/deglaciation
events, resulted in temporal connections with the
Arctic Ocean and changing of the whole salinity
regime in the basin. The Caspian Sea eventually
became isolated and gradually its waters acquired a
unique salt composition, similar to freshwater
but more concentrated. The Black Sea was
freshwater/slightly brackish during the Holocene
but when connected to the Mediterranean Sea
(c.a. 7–6 kyrs) (Dumont 2000; Chepalyga 2007),
the subsequent inflow of seawater caused serious
extinctions of resident biota. The results of this
can be seen in the assemblages of present-day
‘relic’ species now confined to brackish lagoons
and river estuaries. Some 80% of this assemblage
is endemic, including more than one hundred
endemic crustacean species, of which amphipods
are the most prominent group. The long and
complex chain of events which formed the PontoCaspian basin, combined with a harsh continental
climate with large annual fluctuations in water
temperature and dissolved oxygen levels, have
ultimately created an assemblage of euryoecious
and euryhaline species highly tolerant of rapid
environmental change. Such adaptations have
undoubtedly contributed to the success of PontoCaspian invaders during their subsequent invasions
of Central European waters (Bij de Vaate et al.
2002).
In its native range, i.e. the Ponto-Caspian area,
Dikerogammarus villosus inhabits lower courses
of big rivers, such as Danube, Dnieper, Dniester,
Don and Volga as well as coastal lagoons and
limans, however its distribution has not been
studied in details, particularly in the Caspian Sea
area (eg. Carauşu 1943; Carauşu et al. 1955;
Mordukhai-Boltovskoi 1960, 1969; Birstein and
Romanova 1968; Mordukhai-Boltovskoi et al.
1969; Pjatakova and Tarasov 1996) (Figure 3).
Presumably, it is present in most of the major
rivers and limans of the Black, Azov and Caspian
sea basins.
Invasion routes of D. villosus in continental Europe
Jażdżewski (1980) and Bij de Vaate et al. (2002)
distinguished three main aquatic migration
corridors: southern, central and northern, which
provide access routes through Central Europe for
Ponto-Caspian fauna. The southern corridor covers
the Danube and Rhine rivers connected by Main–
Danube Canal. The central corridor constitutes of
the Dnieper and the Pripyat rivers, connected
first to the Baltic Sea basin by the Bug and
Vistula Rivers and then to the North Sea basin
via the Mittelland Canal. The northern corridor
is comprised mainly of the Volga River, the
Beloye, Onega and Ladoga lakes and of the Neva
River that drains to the Baltic Sea. The three
above routes were used by numerous PontoCaspian species (Bij the Vaate et al. 2002; Panov
et al. 2009) to invade the freshwaters of central
and western Europe, however the invasion of D.
villosus presents a unique case (Figure 3).
The D. villosus invasion via the southern
corridor started as early as in 1926, where the
monitoring of macroinvertebrates in the Danube
River first revealed the presence of this species
in Hungary (Nesemann et al. 1995). Then, in the
1950s, the species was detected in Lake Balaton
(Muskó 1989; Muskó 1990; Muskó 1993), shortly
after opening of a canal connecting the lake to
the Danube. The invader then continued its riverine
migration up the Danube, until in 1989 it was
detected in Austria and subsequently, in 1995 in
Slovakia (Šporka 1999). Amongst the main
tributaries of the Danube, an abundant D. villosus
population was detected in the Croatian section
of the River Drava in 2007 (Žganec et al. 2009)
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T. Rewicz et al.
Figure 3. Invasion history and routes of Dikerogammarus villosus spread in Europe. Dates indicate the first record in the particular
section of the invasion route (citations in the text). Red stars indicate known localities within the native range (for details see Supplementary
material - Tables S1 and S2).
Table 1. Presence of Dikerogammarus villosus in the Alpine Lakes.
No
Lake
Country
1
2
3
4
5
6
7
8
9
10
11
12
Leman
Constance / Bodensee
Garda
Neuchatel
Traun
Bienne
Murten = Morat
Zurich
Bourget
Greifen
Zug
Iseo
France, Switzerland
Germany
Italy
Switzerland
Austria
Switzerland
Switzerland
Switzerland
France
Switzerland
Switzerland
Italy
and in the lower section of the River Vah in
Slovakia in 2001 (Brtek 2001, summarised by
Lipták 2013), followed by its spread upstream to
the middle section of the river (Hupało et al.
2014). In 1992, reconstruction of the Ludwig Canal
connecting the Danube, via the Main River, to the
Rhine River was completed (Van der Velde et al.
270
Year of first record of
D. villosus
2002
2003
2003
2003
2003
2005
2006
2006
2007
2008
2010
2011
Source
Bollache 2004
Mürle et al. 2003
Casellato et al. 2006
Lods-Crozet and Reymonnd 2006
Holdich and Pöckl 2007
Lods-Crozet and Reymonnd 2006
Lubini et al. 2006
Steinmann et al. 2006
Grabowski et al. 2007b
Steinmann 2008
Steinmann 2010
Bącela-Spychalska et al. 2013a
2000; Nehering 2002) and D. villosus quickly
penetrated this new waterway, with stable D.
villosus populations being found both in the
upper Rhine (in Bavaria) and the lower Rhine (in
the Netherlands) by 1994 (Bij de Vaate and Klink
1995; Bij de Vaate et al. 2002). Further rapid range
expansion then occurred, with it being detected
Dikerogammarus villosus - profile of the invader
in the River Meuse in 1996 (Josens et al. 2005)
and a year later in the River Saone in France
(Devin et al. 2001; Bollache et al. 2004). Assessment of monitoring covering the years 1997–
2002 showed D. villosus had, by the end of that
period, spread throughout the entire course of the
Rhine, having colonized the Moselle River in
1999, the Seine River in 2000 and the Loire
River in 2001–2003 (Bollache et al. 2004). From
the Rhine River, D. villosus most probably spread
eastwards, entering the central corridor sensu Bij
de Vaate (2002), through the Mittelland Canal to
the River Elbe and thence its largest tributary,
the Vltava (Berezina and Duris 2008). Continuing
this eastward migration, it spread to the Havel,
Spree and Oder Rivers in quick succession, arriving
in the latter by 1999 (Grabow et al. 1998; Zettler
1999; Rudolph 2000). From the Oder River,
D. villosus then spread quickly upstream and
downstream, reaching the oligohaline Szczecin
Lagoon (Jażdżewski et al. 2005; Grabowski et al.
2007c).
Thus far, the final step of this eastward
migration from the southern corridor has been to
the Warta River, the largest tributary of the
Oder. Thus, the species has followed a very long,
circuitous route comprising almost 4500 km in
length, which commenced in the Black Sea
basin, continued through the North Sea basin and
ultimately ended in the Baltic Sea basin (Müller
et al. 2001; Jażdżewski and Konopacka 2002).
Another facet of D. villosus range expansion via
the southern corridor, was colonization of the
Alpine region, with D. villosus being detected in
12 Alpine lakes (Table 1), the River Mincio in
Italy in 2003 (Casellato et al. 2006) and in 2008
it reached Lake Bilancino near Florence (Tricarico
et al. 2010).
Dynamics of the species migration through the
Dnieper River system to Central Europe followed
a scheme, very different from those described
earlier for other invasive species. During the
1950s and 60s the Soviet Union undertook an
extensive dam construction program on the
Dnieper River system, thereby creating several
massive lakes. Ponto-Caspian amphipods, including
D. villosus, were recognized as an important fish
food (Zuravel 1963; Dedju 1967), and deliberately
introduced to these newly established water
bodies to enhance the resident fish food base and
facilitate faster fish production (Zuravel 1965;
Ioffe and Maximova 1968). From there, D. villosus
was able to progress up the Dnieper and then
Pripyat River to the Pripyat-Bug Canal, thus
crossing the Black Sea/Baltic watershed, until in
2003 it was reported from the Bug River, a
tributary of the Vistula River in eastern Poland
(Konopacka 2004). A later record of D. villosus
in the Belarusian part of the Dnieper in 2006
(Mastitsky and Makarevich 2007) confirmed its
spread via the central corridor. After the rapid
spread through the Bug River, D. villosus then
reached the entrance of the Zegrzynski Reservoir
(Grabowski et al. 2007c) and here the range
expansion seemingly stalled for three years until
in 2007, when D. villosus was detected in the
Vistula, the largest river of the Baltic Sea
drainage area (Bącela et al. 2008).
D. villosus has also spread successfully in the
Volga River system, migrating out from the deltaic
system to reach a distance 4000 km upstream by
the middle of the 20th century (MordukhaiBoltovskoi 1960). In 2000 it was detected for the
first time in the Kuybyshev Reservoir, the most
northern record of this species in the northern
corridor (Yakovleva and Yakovlev 2010). However,
surprisingly, this particular range expansion seems
to have stalled in recent years (Yakovleva and
Yakovlev 2010).
Colonisation of Great Britain
The most recent episode in the European range
expansion of D. villosus has seen it escape the
confines of mainland Europe, to reach Great
Britain. Here, in 2010, it was found in the Grafham
Water Reservoir in Cambridgeshire, England
(MacNeil et al. 2010). The source population for
this introduction remains unidentified and requires
a molecular approach to reveal it. This British
introduction provoked a great deal of press interest
and the English and Welsh Environment Agency
set about implementing biosecurity precautions and
new procedures in an attempt to halt its future
spread (Constable and Fielding 2011; Madgwick
and Aldridge 2011). However, despite these efforts,
D. villosus continued to spread rapidly throughout
Britain and within a space of a few months was
reported at Eglwys Nunydd and in Cardiff Bay in
Wales, as well as at several sites in England
(Environmental Agency 2012; MacNeil et al. 2012).
Any which way you can – reasons why
D. villosus has spread so fast
Dikerogammarus villosus possesses several
behavioural traits that probably facilitated its
extremely rapid dispersal throughout Europe.
271
T. Rewicz et al.
Firstly, it has a high tendency to drift in the
water column. Van Riel et al. (2011) showed this
species is a dominant component of the drifting
macroinvertebrate fauna in the River Rhine and
thus may easily colonize rivers by downstream
drift. The species also has a tendency to hide
among zebra mussel Dreissena polymorpha beds
covering the sides of boats and can survive
amongst such mussel bed clusters for at least 6
days even when these boats are out of water
(Martens and Grabow 2008). This ability while
accelerating upstream migration will also, perhaps
more importantly, enable overland transport to
isolated waterbodies. For instance, BącelaSpychalska et al. (2013a) found D. villosus occurs
most frequently in those Alpine lakes experiencing
the highest tourist pressure and presented
experimental evidence suggesting this could be
due to D. villosus being introduced and spread
via sailing ropes and diving equipment.
Is the killer shrimp really a killer?
Dikerogammarus villosus has earned the nickname
‘killer shrimp’, with a body size (30 mm body
length) larger than all other European freshwater
gammarids (Devin et al. 2003), coupled with
massive mouthparts allowing it to overpower and
predate large and well-armored prey (Nesemann
et al. 1995; Mayer et al. 2008; Mayer et al. 2009).
It is a voracious predator, preying upon a wide
range of benthic macroinvertebrates, such as
chironomid, mayfly and dragonfly larvae, aquatic
bugs, leeches, isopods, juvenile crayfish and other
amphipods (Dick et al. 2002; Krisp and Maier
2005; MacNeil and Platvoet 2005; Buric et al.
2009; Platvoet et al. 2009a; Boets et al. 2010;
Hanfling et al. 2011). D. villosus frequently injures
and kills other macroinvertebrate taxa without
consuming them, which additionally increases
this invader’s impacts on prey populations (Dick
et al. 2002). This predation extends to fish eggs
and fry (Casellato et al. 2007; Platvoet et al. 2009b)
and D. villosus can actually function as a top
predator occupying the same trophic level as fish
and the largest predatory macroinvertebrates
(van Riel et al. 2006). However D. villosus could
actually have a greater impact than fish predators
on resident benthic prey communities, because it
occurs at higher abundances and has a body size
similar to potential prey. This latter factor allows
D. villosus to penetrate small refugia, so there is
no effective hiding place for prey taxa, as there
would be from fish predators (MacNeil et al.
272
2011). The mere presence of D. villosus has also
been shown to lead other amphipods to leave
previously occupied benthic refugia, swim up in
the water column and become more vulnerable to
fish predators (Kinzler and Maier 2006).
D. villosus is a voracious predator so obviously
it cannot be classified as predominantly a shredder
of leaf litter, as are most native amphipods (Mayer
et al. 2009; MacNeil et al. 2011). However, it is not
a strict predator, feeding as it does on detritus,
carrion, and even microalgae (Dick et al. 2002;
Kley and Maier 2005; Platvoet et al. 2006a; Mayer
et al. 2008). Indeed, juveniles are very efficient
consumers of plant material (micro-algae), before
they reach maturity and become extreme
opportunists (Platvoet et al. 2006a; Maazouzi et
al. 2007; Mayer et al. 2008). This is confirmed by
ultrastructure studies of its mouthparts, which show
a lack of morphological specialization to eat any
particular food type (Platvoet et al. 2006a; Mayer et
al. 2012). This capability to function mainly as a
major predator but also act as an omnivore when
other food resources are more plentiful or easier
to obtain, confers a huge competitive advantage
over many other amphipod species resident in
European freshwaters. An indirect effect of D.
villosus invasion may also be disruption of leaflitter processing and shredder efficiency, as
laboratory studies have shown that predation by
and even the mere presence of, D. villosus, can
curtail the activity of macroinvertebrate taxa
engaged in leaf shredding (MacNeil et al. 2011).
This could have profound consequences on energy
cycling in the invaded ecosystem as leaf litter
may cease to be broken down and so remain
unavailable to the rest of the resident community
who rely on the breakdown products of litter
‘shredding’ (MacNeil et al. 1997; 2011).
It is unsurprising that the invasion and spread
of D. villosus, with its high population densities
and predatory disposition, has generated many
interactions with resident amphipod species, both
natives and previously successful invaders (Dick
and Platvoet 2000; Dick et al. 2002; Kinzler and
Maier 2003; Kley and Maier 2005; Platvoet et al.
2006a; van Riel et al. 2009). One such example
is the elimination of the native Gammarus duebeni
Liljeborg, 1852 from the Ijsselmeer/Markermeer
lake in the Netherlands. This lake was previously
inhabited by G. duebeni coexisting with a North
American invader, Gammarus tigrinus Sexton,
1939. Both these species inhabited basalt boulders
lining the shore-line and there was direct
competition between newly arrived D. villosus
Dikerogammarus villosus - profile of the invader
with these species for this habitat. Subsequent
monitoring revealed D. villosus had replaced the
other species in the boulder zone, with G.
duebeni completely disappearing and G. tigrinus
retreating to softer sediments, deeper in the lake
(Dick and Platvoet 2000; Dick et al. 2002; Platvoet
et al. 2006b). The impact of D. villosus on resident
amphipod assemblages of the River Rhine has also
been drastic. After its arrival, the abundance of
the resident Chelicorophium curvispinum (G.O.
Sars, 1895) was greatly reduced (van Riel et al.
2006) and G. tigrinus was displaced from its
preferred stony habitat to less favorable ones
(van Riel et al. 2006). Such patterns of displacement
tend to be repeated in all newly colonized areas
and niche partitioning presents itself as the only
mechanism allowing resident amphipods to coexist with D. villosus, being confined to
macrophytes, weeds or soft sediments, while
being eliminated in stony habitats by D. villosus
(Devin et al. 2003; Kley and Maier 2005; van Riel
et al. 2007; Felten et al. 2008b; Hesselschwerdt
et al. 2008; MacNeil et al. 2008; Kley et al. 2009;
Boets et al. 2013a).
Intraguild predation (IGP) or predation between
competitors belonging to the same ecological
guild (Polis et al. 1989), despite being considered
an unstable phenomenon (e.g. Holt and Polis 1997),
is a widespread interaction in natural food webs
(Arim and Marquet 2004) and is increasingly
acknowledged as a major driver of rapid species
exclusions during biological invasions (e.g.
Snyder et al. 2004; Wang et al. 2013). IGP by D.
villosus of several amphipod species including
G. duebeni, G. tigrinus, Gammarus fossarum
Koch, 1836 and Gammarus roeselii Gervais, 1835
has been witnessed in laboratory studies (Dick
and Platvoet 2000; Kinzler and Maier 2003). IGP
of Gammarus spp. by D. villosus occurs on both
newly moulted (Kinzler and Maier 2003) and
intermoult (Dick and Platvoet 2000) individuals,
albeit less frequently on the latter. Although
Kinzler et al. (2009) found no superior IGP by
D. villosus upon other similar sized Dikerogammarus species, such results appear counter
intuitive when field observations indicate
displacement of species such as D. haemobaphes
by D. villosus incursions (Grabowski et al.
2007c; Kinzler et al. 2009).
One important factor facilitating D. villosus
colonization is the presence of the zebra mussel
Dreissena polymorpha. This globally widespread
bivalve lives in colonies and these shell ‘beds’
provide the heterogeneous, hard structured habitat
ideal for D. villosus (Devin et al. 2003). D. villosus
also feeds on the zebra mussel’s byssus threads
(Platvoet et al. 2009b), faeces and pseudofaeces
(Gergs and Rothhaupt 2008a). The biomass
accumulating in zebra mussel colonies also forms
a perfect food base for chironomid larvae, a major
D. villosus prey item (Maier et al. 2011) and such
very abundant and highly calorific food items
promotes the rapid growth and development of
D. villosus (Gergs and Rothhaupt 2008b). Laboratory experiments have shown D. villosus grows at
double the rate when consuming chironomid
larvae, rather than biodeposited material or
conditioned leaves (Gergs and Rothhaupt 2008a).
In comparison, the growth rate of G. roeselii is
half that of D. villosus, when provided with the
same number of larvae (Maier et al. 2011).
An approach to assessing the ecological impact
of an invading species such as D. villosus on a
resident community is the comparison of its resource
uptake rate or predatory ‘capacity’ compared with
that of a trophically analogous resident species
(Dick et al. 2013; Dodd et al. 2013). This predatory
‘capacity’ can be quantified by measuring the
relationship between resource consumption rate
(in this case predation rate) and resource density
(in this case prey availability) in a ‘functional
response’ (Abrams 1990). A Type II functional
response represents a consumption rate which
increases with prey density but then declines to
an asymptote as prey handling time becomes a
limiting factor (Holling 1966). Dodd et al. (2013)
compared the functional responses of D. villosus
to that of G. pulex Linnaeus, 1758, in respect of
three common prey, Asellus aquaticus (Linnaeus,
1758), Chironomus sp. and Daphnia magna
Straus, 1820. Both large D. villosus individuals
and those matched for body size with G. pulex,
showed higher Type II functional responses than
G. pulex in respect of two prey types and similar
for the third. Thus, D. villosus showed higher
maximum feeding rates than G. pulex on both
A. aquaticus and D. magna (similar for Chironomus
sp.), making it a more efficient predator and
consequently likely to have a greater impact on
prey populations. In addition, mixed prey type
experiments showed that D. villosus was significantly more indiscriminate in prey selection than
G. pulex and this may be crucial as the ecological
impact of an invader possessing indiscriminate
feeding habits is likely to be far greater than a
more selective one.
273
T. Rewicz et al.
As quick and as many as possible
Biological invasion is a combination of stages
and barriers that the future invader has to cross
(Blackburn et al. 2011). Thus only the species that
can be transported, introduced, survive, reproduce
and spread may invade new territories successfully.
Several biological traits promote invasion of
these species and in general species with greater
dispersal ability, ecological generalization and
greater reproductive rate should be more likely to
colonize new areas. Thus detailed knowledge on
the life history of an invader is crucial for
estimating the likely invasion success of the invader
(Olden et al. 2006). Dikerogammarus villosus
seems to be a model, almost ‘perfect’ invader in
these aspects. Its life cycle and reproductive
behaviour has been investigated in both its native
and invaded range (Mordukhai-Boltovskoi 1949;
Kley and Maier 2003; Piscart et al. 2003; Devin
et al. 2004; Kley and Maier 2006; Pöckl 2007;
Pöckl 2009). Grabowski et al. (2007a) summarised
all available data on the life history traits of
amphipods native and invasive to Central Europe,
including D. villosus, and found that, generally,
alien species were characterised by larger brood
sizes, higher partial fecundity, earlier maturation
and a higher number of generations per year,
than native species.
Dikerogammarus villosus grows faster than
many freshwater amphipods (Piscart et al. 2003)
and reaches sexual maturity earlier, with females
as small as 6 mm in length having broods
(Mordukhai-Boltovskoi 1949; Piscart et al. 2003;
Devin et al. 2004; Pöckl 2007; Pöckl 2009). This
size is achieved between the 33rd and 60th day
of life, depending on temperature (Piscart et al.
2003; Pöckl 2009). The results from these latter
studies contrast with Mordukhai-Boltovskoi (1949)
who reported D. villosus taking 110 days to
achieve sexual maturity but this latter study did
not specify the temperature or body size achieved
by this time. However, compared to D. villosus,
other European amphipods lag far behind in the
time needed to attain sexual maturity, for example
Gammarus pulex needs 133 days at 15°C (Welton
and Clarke 1980), Gammarus fossarum 96 days
and Gammarus roeselii 85 days at 20.2°C (Pöckl
1992). Once sexual maturity has been reached,
the breeding period of D. villosus is also
relatively long and under European climatic
conditions, ranges from 9 to 12 months (Ciolpan
1987; Devin et al. 2004; Pöckl 2007; Pöckl 2009)
which contrasts with 4 to 9 months for most
other amphipod species inhabiting the same
274
geographic region (summarized in Grabowski et al.
(2007a) (Table 2).
In summary, these life history traits make this
species an excellent colonizer, with one large
female D. villosus capable of producing more
offspring in one brood than a female G. fossarum,
G. pulex, or G. roeselii could do during their
entire lives. This means, potentially, it would
require only one or two D. villosus females to
establish a viable population in a newly
colonized water-body (Pöckl 2007).
The killer shrimp hitchhikers
Several “hitchhikers” in the form of parasites,
probably accompanied Dikerogammarus villosus
during the invasion process and so were spread
with their hosts into new and perhaps naive systems.
A survey investigating parasite diversity in the
central corridor showed that Ponto-Caspian
amphipod hosts have a significant role as vectors
for gregarines which belong to protozoa and for
microsporidia (Fungi) (Ovcharenko et al. 2008;
Ovcharenko et al. 2009). It appeared that D. villosus
is a host for these parasite taxa, with four gregarine
species identified infecting D. villosus, including
one acanthocepahalan and several microsporidians
(Table 3). On the other hand a thorough survey
by Bojko et al. (2013) focusing on populations of
the killer shrimp that colonised UK and two
continental populations from the invaded range
confirmed “the parasite release hypothesis” in
case of the UK populations. They showed significantly lower diversity and prevalence of parasites
(especially microsporidians were not noticed at
all) compared to the continental populations.
Microsporidia have been well studied as
obligatory intracellular parasites, infecting many
animal taxa and are very common amphipod
parasites. They are transmitted both horizontally
and vertically (MacNeil et al. 2003; Haine et al.
2004; Haine et al. 2007) and depending on the
transmission mode they can be either lethal or
relatively harmless to their hosts (Terry et al.
2004). While most microsporidian species are
relatively rare in D. villosus host populations,
with prevalence typically below 4 % (Wattier et
al. 2007), Cucumispora dikerogammari which
infects D. villosus within and outside its native
range, can attain a prevalence of up to 74%
(Wattier et al. 2007; Ovcharenko et al. 2010; BącelaSpychalska et al. 2012). This microsporidian
parasite was seldom detected in amphipod hosts
other than D. villosus and when it was, its host range
was restricted to other Ponto-Caspian amphipods
Dikerogammarus villosus - profile of the invader
Table 2. Data matrix of life history traits and ecological tolerance of Dikerogammarus villosus and other gammarid species (after Grabowski
et al. 2007a, modified). x – the number of generation per year cannot be estimated as the reproduction is continuous throughout the year.
Species
mean
breeding
female
size (mm)
mean
partial
brood
fecundity
size
breeding
period
in
months
maturity
index
number of
generations per
year
salinity
tolerance
human
impacts
G .fossarum
10.14
16.88
1.66
10.00
0.79
2
1
1
G. lacustris
G.varsoviensis
G. pulex
G.leopoliensis
G. balcanicus
G. roeselii
11.28
13.50
8.90
9.40
9.10
12.55
18.75
25.17
14.79
16.70
7.88
25.60
1.66
1.86
1.66
1.78
0.87
2.04
4.00
5.00
10.67
7.00
7.00
6.00
0.71
0.74
0.88
0.79
0.84
0.68
1
1
1
1
1
2
2
1
2
1
2
1
2
2
2
1
2
2
P. robustoides
12.65
64.45
5.10
7.00
0.63
3
3
3
D. haemobaphes
10.99
42.84
3.90
5.50
0.57
3
2
3
D. villosus
11.39
50.66
4.45
11.00
0.57
3
2
3
O. crassus
8.81
25.33
2.87
7.00
0.68
3
3
2
C. ischnus
7.83
17.33
2.21
8.00
0.64
2
2
2
G. tigrinus
7.94
20.31
2.56
9.00
0.50
3
3
3
L. scutariensis
E. cari
7.90
5.60
15.53
8.6
0.51
1.59
12.00
10.00
0.76
-
x
x
1
1
1
1
Sources (combined with own data)
Jazdzewski 1975; BrzezinskaBlaszczyk and Jazdzewski 1980
Hynes 1955; Hynes and Harper 1992
Jazdzewski 1975; Konopacka 1988
Hynes 1955; Jazdzewski 1975
Zielinski 1998
Jazdzewski 1975; Zielinski 1995
Jazdzewski 1975; own data
Bacela and Konopacka 2005; Dedju
1966, 1967, 1980; Kasymov 1960;
Musko 1993; Kiticyna 1980;
Kurandina 1975
Devin et al. 2004; Kley and Maier
2003; Mordukhai-Boltovskoi, 1949
Kurandina 1975
Kley and Maier 2003; Konopacka
and Jesionowska 1995; Kurandina
1975; Mordukhai-Boltovskoi,1949
Bousfield 1958; Chambers 1977;
Pinkster et al. 1977 ; Steele and
Steele 1975
Grabowski et al. 2014
Zganec et al. 2011
Table 3. Known parasites infecting Dikerogammarus villosus.
Observed
max
prevalence
Parasite
Trematoda
Plagioporus skrjabini
Not identified
nd
<2%
Acanthocephala
Pomphorhynchus tereticollis (Rudolphi, 1809)
0.04%
Apicomplexa,
Gregarinia
Cephaloidophora sp.
65%
Cephaloidophora similis Codreanu-Balcescu, 1995
na
Microsporidia
Bacteria
Viruses
Cephaloidophora mucronata Codreanu-Balcescu,
1995
Uradiophora sp.
Uradiophora longissima (von Siebold in von
Kölliker, 1848)
65%
Uradiophora ramosa Balcescu-Codreanu, 1974
na
na
na
Cucumispora = Nosema dikerogammari (Ovcharenko
and Kurandina, 1987)
74%
Nosema granulosis Terry et al. 1999
4%
Dictyocoela muelleri
D. berillonum
D. roeselum
not identified
Dikerogammarus villosus bacilliform virus
3.4%
2%
2%
< 1%
< 1%
Geographic region
Source
the Volga River,
UK fresh waters
the River Rhine, Vistula
River
Chernogorenko et al. 1978
Bojko et al. 2013
Emde et al. 2012; Bojko et al.
2013
UK fresh waters
Bojko et al. 2013
invaded waterbodies: i.e.
Vistula, Oder
invaded waterbodies: i.e.
Vistula, Oder
UK fresh waters
invaded waterbodies: i.e.
Vistula, Oder
invaded waterbodies: i.e.
Vistula, Oder
Bojko et al. 2013
whole range, despite UK
Ovcharenko and Kurandina,
1987; Wattier et al. 2007;
Ovcharenko et al. 2009;
Bącela-Spychalska et al. 2012;
Bojko et al. 2013
invaded waterbodies i.e. the
upper Danube, the Rhine,
Seine, Loire
Rhine drainage
Meuse River
the upper Danube, the Rhine
the Vistula River
the Vistula River
Ovcharenko et al. 2009
Ovcharenko et al. 2009
Ovcharenko et al. 2009
Ovcharenko et al. 2009
Wattier et al. 2007
Wattier et al. 2007
Wattier et al. 2007
Wattier et al. 2007
Bojko et al. 2013
Bojko et al. 2013
275
T. Rewicz et al.
such as D. haemobaphes, Echinogammarus ischnus
and Chelicorophium curvispinum and always at a
low prevalence (less than 4%). Such findings
indicate that C. dikerogammari is virtually specific
for the D. villosus host (Bącela -Spychalska et al.
2012). This parasite is virulent only in the later
stages of infection, with only symptomatic
individuals with a high parasite load exhibiting
increased mortality; with infected but asymptomatic
ones exhibiting the same survivorship as uninfected
individuals (see Bącela-Spychalska et al. (2012).
Transmission to other non-D. villosus hosts is
also limited by a lack of macroinvertebrate predation
on D. villosus. In summary, this particular microsporidian cannot, as yet, be considered as a new
threat to resident amphipods in systems invaded
by D. villosus. However, the parasite may modify
the impact of D. villosus on the macroinvertebrate
assemblages within invaded systems, as it
significantly diminishes predation rate by the D.
villosus host, thus potentially reducing predation
pressure (Bącela-Spychalska et al. 2013b).
The killer shrimp as a model species for
Amphipod morphology – AMPIS
The Amphipoda Pilot Species Project (AMPIS),
based on complex images in the macro- and microscanning electron microscope (SEM) scale, was
initiated in 2005 (Platvoet et al. 2006b) to address
the lack of a comprehensive database of amphipod
morphology. The plan was to create a standardized
database containing complete descriptions of several
species of Amphipoda, which could be used as
templates for taxonomic and morphological
studies. Dikerogammarus villosus was chosen as
a pilot species for the program, due to its
prominence as an important European invader,
its large body size, predatory behaviour and its
potential impact on invaded communities
(Platvoet et al. 2007).
Extensive study of the body surface ultrastructure
of D. villosus revealed the presence of some
unexpected structures. For instance females have
two pores on the first pereionite, which are larger
than any other pores in the amphipod cuticle.
Although the function of such structures remains
unidentified, it has been hypothesized that they
are associated with reproductive behavior and
the release of chemical attractants (Platvoet et al.
2006b). Other structures described may be part of
the ‘locking-on’ system used in amplexus stage,
as males and females pair up. For instance,
276
females possess swollen edges of the first and
fifth pereionite, which fit or ‘lock’ to the male’s
first gnathopod’s palm and palmar angle. This
would enable correct size-selection during mate
choice and it could have other functions such as
stimulating hormonal processes in males and
females, stabilization of amplexus and promote
energy conservation during paired swimming
(Platvoet et al. 2006c). SEM pictures of the D.
villosus cephalon reveal depressions on each side,
probably signifying the presence of statocysts or
balance sensory receptors. These are associated
with geosense and spatial orientation and in
D. villosus may act as movement detectors and/or
monitors of hydrostatic pressure (Platvoet et al.
2006d). In common with many other amphipods,
D. villosus possesses a lateral line organ comprised
of two rows of specialized receptors units on
each side of the body. Similar to other animals
(particularly fish and amphibians) such receptors
may be linked to chemo-, mechano-, and electrosensory functions (Platvoet et al. 2007).
How to kill the killer shrimp – the ‘perfect’
invader has weaknesses
No invader is perfect, not even Dikerogammarus
villosus and any perceived weaknesses need to
be exploited if this invader’s spread and impact
are to be minimised. In its native range, D. villosus
occurs in many different types of water bodies
including limans, lakes, reservoirs and the mouths
and main channels of large rivers and many of
these exhibit very changeable physico-chemical
conditions, particularly in respect of salinity and
temperature regimes. D. villosus has been found
surviving in brackish water up to 10 psu, and can
even acclimatize to 20 psu under laboratory
conditions (Bruijs et al. 2001; Brooks et al. 2008).
In addition, although degradation of eggs has
been observed at 20 psu, at salinities as high as
15 psu development of embryos is still possible
and hatching of juveniles has been recorded
(Bącela-Spychalska pers. obs). This ability
significantly increases its potential for long
distance transport within ship ballast water and
will consequently enhance range expansion (Brooks
et al. 2008; Santagata et al. 2009; Piscart et al.
2011). However, this tolerance does have its
limits, so a simple but effective tool against
introduction of D. villosus via overseas shipping,
would be to replace brackish ballast water obtained
from ports/harbours with high salinity open ocean
water (Santagata et al. 2008; Santagata et al. 2009).
Dikerogammarus villosus - profile of the invader
We have previously noted the ability of
D. villosus to survive at least 6 days outside a
waterbody and within damp zebra mussel shell
clusters covering the sides of boats (Martens and
Grabow 2008). In addition, Bącela-Spychalska et
al. (2013a) has reported survival for 3–5 days
within the folds of a moist neoprene diving suit.
D. villosus has also been observed to survive up
to 6 days in a pile of macrophytes and roots left
out of water (Rewicz et al. pers. obs.). Poznańska et
al. (2013) also reported that individual D.
villosus exposed to air without shelter exhibited
grouping behaviour, enhancing their survival to
desiccation as compared to more exposed single
individuals who would be more liable to drying
out. This high tolerance to air exposure obviously
greatly increases the potential for overland transport
and rapid range expansion. Despite this, temperature
tolerance of D. villosus is similar to the majority
of native European freshwater amphipods, with a
critical threshold level of 31°C, a temperature
lower than some other invaders, such as Gammarus
tigrinus (37°C) or Echinogammarus ischnus
(35°C) (Wijnhoven et al. 2003; van der Velde et
al. 2009). However, Maazouzi et al. (2011) reported
that under laboratory conditions the limiting
temperature for D. villosus was as low as 26°C
compared to 30°C for the native Gammarus pulex
and the authors accounted for this difference to
the earlier studies by the relatively long duration
of their experiment (15 days). Our own data
(unpublished) obtained from the native range of
D. villosus, shows that water temperatures ranged
from 25°C to almost 29°C in July. Wijnhoven et
al. (2003) showed that D. villosus has a reduced
tolerance to temperature in waters with low
conductivity and this could account for its
preference for larger rivers with higher temperatures
and conductivities, compared to smaller tributaries/
streams with lower temperatures/conductivities
(Grabowski et al. 2009).
Amphipods are sensitive to a wide range of
toxicants (Felten et al. 2008a) and are increasingly
recognised as important bioindicators in ecotoxicological tests (Kunz et al. 2010). D. villosus, being
an increasingly common species in European
watercourses, has also started to be recognized as
a robust pollution indicator (Sebesvari et al. 2005).
It has been used to measure metal bioaccumulation
and provide information about contamination levels
in aquatic ecosystems (Barkács et al. 2002). For
instance, Sebesvari et al. (2005) showed that D.
villosus is a useful bioindicator of tin, as its
concentration in the amphipod’s tissues has a
strong correlation with its background environmental concentration. Similarly, D. villosus has a
physiological ability to respond to higher copper
concentrations by decreasing its total fatty acids
content (Maazouzi et al. 2008; Sroda and CossuLeguille 2011). Interestingly, other amphipods such
as Gammarus roeselii are more sensitive to copper
levels, so copper pollution could further enhance
D. villosus invasion by weakening/eliminating
potentially competitive native species. In contrast,
D. villosus is very sensitive to fluoride (Gonzalo
et al. 2010) and cadmium (Boets et al. 2012), and
accumulates these quicker than many resident
amphipods. Thus, high fluoride/cadmium levels
may make some water bodies relatively resistant
to D. villosus invasion and successful establishment.
In addition, cadmium exposure has been shown
to interfere with antipredatory behavior (i.e.
aggregation with conspecifics, refuge use,
exploration and mobility) in D. villosus and may
cause disrupt function of chemosensors (Sornom
et al. 2012). Such high sensitivity for various
chemical stressors displayed by D. villosus
populations could be related to the rapid expansions
into the new areas, reflecting low genetic diversity
in founder populations and a bottle-neck effect
(Piscart et al. 2011; Boets et al. 2012).
Establishing methods to both eradicate
Dikerogammarus villosus and prevent further
spreading are current priorities of government
environmental protection agencies such as the
Environment Agency in England and Wales.
High attachment abilities of D. villosus to objects
submerged in the water like ropes, wet suits,
boat hulls, nets etc. has already been recognized
(Bącela-Spychalska et al. 2013a) and there is an
increasing focus on ways to ‘stop the spread’ by
concentrating on ways to sterilize such equipment
of any potential D. villosus ‘hitchhikers’. Such
research has been conducted by the Centre for
Environment, Fisheries and Aquaculture Science
(CEFAS) and the Science and Technical Advisory
Group (STAG) in Great Britain. Thirteen chemical
and physio-chemical treatments were assessed as
potential D. villosus eradaticators i.e.: pH, salinity,
iodine/iodophor, chlorine/sodium hypochlorite,
virkon S, temperature, acetic acid, methanol,
citric acid, urea, hydrogen peroxide, carbonated
water and sucrose (Stebbing et al. 2011). After
considering all these options, the most effective
eradication method was found to be simple
application of heated water (50°C), which resulted
in 100% mortality level instantly. In contrast,
carbonated water only induced narcosis, but is
277
T. Rewicz et al.
cheap and easier to implement in the field. It is
hoped public education campaigns like the "check,
clean, dry" (GB non native species secretariat)
approach in the U.K., coupled with establishing
decontamination places and protocols in marinas,
reservoirs and other popular tourist areas, will
stop or at least slow down the spread of
D. villosus.
The recent arrival of D. villosus in British
freshwaters
Since the initial detection of Dikerogammarus
villosus in British freshwaters in 2010, its
subsequent range expansion and that of another
recent Ponto-Caspian amphipod invader to Great
Britain, Dikerogammarus haemobaphes is currently
the focus of much study (Gallardo et al. 2012;
MacNeil et al. 2012; Gallardo and Aldridge 2013a;
Gallardo and Aldridge 2013c; MacNeil et al. 2013).
Spread of both invaders has been relatively rapid
(MacNeil et al. 2013), leading to the instigation
of killer shrimp monitoring programmes and
evaluation of risk factors that could facilitate/
accelerate range expansion (MacNeil et al. 2012;
2013). Based on climatic (e.g. temperature),
physico-chemical (level of oxygen, conductivity)
and socio-economic factors (human activities)
SDM (species distribution) and HSM (habitat
suitability) models were developed (Gallardo et
al. 2012; Boets et al. 2013b; Gallardo and
Aldridge 2013a; 2013b; 2013c). These multifactor
models assess the invasive adaptations and
‘potential’ of D. villosus and predict those areas
most vulnerable to D. villosus invasion.
Geographically, about 60% of Great Britain was
found to be potentially suitable for D. villosus,
including the vast majority of central and
southern England, with areas containing harbours,
ports and lakes with high angler/tourism pressure
being particularly vulnerable to invasion (Gallardo
et al. 2012; Gallardo and Aldridge 2013c). In
addition, MacNeil and Platvoet (2013) highlighted
the fact that artificial in-stream structures such as
fish passes and bank reinforcements may
represent optimal habitat types for D. villosus
and D. haemobaphes. Indeed, the presence of
such structures may allow these invader species
to penetrate small rivers and facilitate invasion
of in watercourses which were previously
considered unsuitable for these species in terms
of ‘natural’ habitat with muddy, soft substrates.
In the long term, climate change may promote
the establishment of new invaders and facilitate
278
further spread of existing invaders such as D.
villosus and D. haemobaphes, as invaders may
be better adapted to cope with changing conditions
than native species (Gallardo and Aldridge 2013b;
2013c).
Lessons for the future
Dikerogammarus villosus has earned its moniker
of the ‘killer shrimp’. It is a voracious predator
and a very successful invader, capable of rapid
range expansion, is highly adaptable to new
environmental regimes and physiologically tolerant
enough to survive transport in both ship ballast
water and overland transport. Undoubtedly, it can
have profound impacts on resident macroinvertebrate communities and its arrival has negative
connotations for biodiversity. The impacts of this
species can be so significant to the structure of
invaded communities and consequent functioning
of ecosystems, that we can propose that in many
invaded systems, it effectively acts as an
‘ecosystem engineer’ (van Riel et al. 2006).
When considering the potential negative impacts
of D. villosus invasion and range expansion in
Europe and whether it is worthwhile to expend
large amounts of resources in an attempt to stop
or even slow the spread of this invader, we must
first acknowledge that in common with many
other invasion scenarios, anthropogenic pressures
fuelled this particular invasion (Pyšek et al.
2010). Construction of canals, alteration of river
flow regimes, in-stream engineering altering
substrate types and bankside structures, increasing
industrial pollution and last but not least, the
enormous traffic flow of boats and barges
throughout Europe’s watercourse have all
contributed to the rapid spread and successful
establishment of this species. Although it is an
unpalatable truth, in all likelihood the ecosystem
changes wrought by D. villosus invasion are
profound and probably irreversible. Recently
implemented restoration programmes for European
rivers, may make them less suitable for D.
villosus, but this is mere speculation at present.
Realistically it will be impossible to eliminate D.
villosus from invaded European rivers or prevent
its further spread to interconnected river
networks. Although this seems desperate, we can
take some positive steps. We need to implement
D. villosus monitoring systems, to provide crucial
information on spread, vectors, biology, impact
on local biota and subsequent economic impacts.
Measures should be implemented to stop
Dikerogammarus villosus - profile of the invader
overland transport of this species to isolated
river systems or lakes. Thus, acknowledging the
biosecurity risk posed by people using different
waterbodies for recreational purposes, specific
procedures exist to stop the accidental spread of
invasive species. For example, in Great Britain,
cleaning boat sides and propellers as well as
sport gear has been recommended as a standard
procedure, before leaving an invaded or invasion
risk site or moving onto a new site (Madgwick
and Aldridge 2011; Anderson et al. 2014). Such
boat cleaning procedures have already been
instigated as standard practice at Grafham Water
Reservoir (MacNeil, pers. obs.). The adoption of
such rigorous biosecurity measures may be
crucial in the protection of the unique freshwater
ecosystems of southern Europe with their great
biodiversity and relatively pristine macroinvertebrate and fish communities. Indeed, catastrophic
impacts have already been witnessed in Lake
Garda, where the killer shrimp has decimated the
population of the resident amphipod Echinogammarus stammeri (S. Karaman, 1931), as well
as preying on the eggs and fry of native fish
(Casellato et al. 2006; Ciutti et al. 2011). The
ancient Lake Ohrid in the Balkan Peninsula is an
example of a water body where D. villosus could
have similar devastating impacts. We regard this
as particularly vulnerable as at least 34% of the
resident animal assemblage is endemic, giving it
the highest endemic diversity among all the
ancient lakes in the world taking into account the
lake surface area (Albrecht and Wilke 2009). In
some animal groups, such as amphipods, more
than 90% are endemic to this lake (Wysocka et
al. 2013). The arrival of a highly competitive and
predatory invader such as D. villosus has the
potential to severely reduce this rich native
diversity and this threat is unfortunately growing
as tourist develop-ment rapidly increases, as this
picturesque and rather isolated lake becomes an
increasingly attractive destination for boating
and diving – activities already proven to
facilitate spread of the invader. Similar risk
factors apply to other Balkan ancient lakes such
as the Shkoder, Prespa, Trichonis and Doiran.
The biodiversity of these lakes has not been well
studied so the arrival of D. villosus could
irreversibly change these lake assemblages, even
before they have been documented pre-invader
impact.
It should be acknowledged, that in practical
terms, we consider it impossible to eradicate this
invader or effectively stop its expansion within
invaded European freshwaters. Thus it worthwhile
to consider an alternative strategy of focusing
efforts/resources on preventing the spread of the
killer shrimp to isolated basins, such as Alpine
lakes or areas of the defined or putative high
freshwater endemic diversity. Taking into account
the isolation of the lakes from the river systems
already invaded, the previously discussed
preventative measures to stop spread by boats and
diving equipment, if applied vigorously, should
greatly reduce the risk of D. villosus invasion.
Another potential destination for D. villosus is
the system of the North American Laurentian
Great Lakes (U.S. EPA 2008). This has already
been invaded by several Ponto-Caspian species
conveyed in ship ballast water (i.e. zebra mussel
(Dreissena polymorpha) and quagga mussel
(D. bugensis), fishhook waterflea (Cercopagis
pengoi), gammarid Echinogammaraus ischnus
and gobies round boby (Neogobius melanostomus)
and tubenose goby (Proterorhinus semilunaris)
(Ricciardi and MacIsaac 2000). Again, simple
measures, if commonly applied, may prevent
such introductions. However, there is still much
to be done to improve the tools against
introduction of invasive species in the Great
Lakes. In case of ballast water management, only
ballast water exchange and saltwater flushing are
mandatory till the year 2016 (Government of
Canada 2006; SLSDC 2008). After that time the
management should be improved by implementing
new, and hopefully, more efficient systems to
avoid exchange of biota (Wang et al. 2012;
Briski et al. 2013). It is also stressed that
preventing non-native species spread in the
Northern America should be also implemented
for the inland ship transport as the intensive
water exchange via ballast tanks is present
between Saint Lawrence River to the Great
Lakes (Adebayo et al. 2014). However, till now
there is “easy entrance” for the killer shrimp to
invade the North American Lakes as the
preventing method used now are not efficient for
Ponto-Caspian invaders as is D. villosus (U.S.
EPA 2008).
The killer shrimp is a prime example of an
invader, whose spread has focussed the attention
of the international scientific community and in
particular government agencies tasked with
nature conservation and the protection of aquatic
ecosystems (GB NNSS 2011). This interest has
been generated by the relatively early alerts on
the potentially highly deleterious impacts the
invader posed for vulnerable benthic communities.
279
T. Rewicz et al.
Given this level of interest in both the scientific
and general media, it remains surprising, that
although over the past two decades, many aspects
of D. villosus ecology have been relatively well
studied, studies on the factors either facilitating
or mitigating against its spread remain scarce
(GB NNSS 2011; Stebbing et al. 2011). This
knowledge ‘gap’ undoubtedly has contributed to
the lack of preventive measures that could be
practically undertaken to stop or at least slow
down its spread that can be estimated using risk
assessments. Only recently have such measures
been tentatively proposed in places such as Great
Britain, and only after a multitude of rivers,
lakes and reservoirs have already been invaded
(GB NNSS 2010; Madgwick and Aldridge 2011).
In hindsight, the case of the killer shrimp may
provide a general lesson for invasion ecologists,
that studies on the underlying mechanisms of
invader spread should be undertaken at earliest
stage of the invasion process. Only then can one
hope to slow its progress within invaded systems
and more importantly prevent it reaching new
systems, before it causes irreversible ecosystem
changes. A new EC proposal on the prevention
and management of the introduction and spread
of invasive alien species was agreed in March
2014 [COD(2013)0307].This new regulation
applies a list of invasive species of “Union
concern” that should not be introduced, transported,
placed on the market, offered, kept, grown or
released into the environment. In the case of
invasive species already introduced, the Member
States will be responsible for establishing the
method of elimination of the species from the
environment. It is not known, at this time,
whether the killer shrimp, will be listed as a
species of “Union concern”. In a very thought
provoking review, Richardson and Ricciardi (2013)
pointed out that despite its critics, invasion
ecology remains a thriving and increasingly
relevant science, especially to anyone concerned
with preventing the loss of biodiversity and
ecosystem services. To conclude on a relatively
positive note, the invasion and spread of the
‘killer shrimp’ has undoubtedly led to
significantly increased financial support and
political/public interest for research into alien
species, the mechanisms of invasion and potential
methods to prevent the spread of damaging
invaders and their associated negative impacts on
taxonomically, culturally and economically valuable
ecosystems. In the end, the least damaging invasion
is the one which is prevented from happening.
280
Conclusions
A. Dikerogammarus villosus has earned its
moniker of the ‘killer shrimp’ and although not
‘perfect’, it is a very, very successful invader. It
is a voracious predator capable of having a
profound impact on freshwater macroinvertebrate
community structure and function. It is highly
adaptable, physiologically tolerant and its
continued rapid range expansion has negative
connotations for native biodiversity on both a
European and potentially global scale.
B. This Ponto-Caspian amphipod has spread
within a few decades throughout Europe’s inland
waters. Its range extension has been associated
mainly with commercial shipping in large
waterways (i.e. the Danube, Rhine, Mittelland
Canal – a circuitous route comprising almost
4500 km) but it has also been transported
overland to many isolated Alpine lakes and has
recently moved out of mainland Europe to reach
the British Isles. Further potential destinations
for D. villosus include the North American
Laurentian Great Lakes system.
C. Inherent life history traits make D. villosus
an excellent colonizer, with one female capable
of producing more offspring per brood than
females of most native European Amphipod
species. Potentially, it would require only one or
two D. villosus females to establish a viable
population in a newly colonized water-body.
D. The capability to function as both a highly
efficient predator and also as an extreme
opportunist omnivore as the need arises confers a
huge competitive advantage over many other
European macroinvertebrate taxa, including
previously successful invaders.
E. The invasion of isolated Alpine Lakes by
D. villosus has been linked with overland transport
associated with recreational activities. Similar
transport risk factors also apply to several Balkan
ancient lakes. The biodiversity of these latter
lakes has not been well studied so the arrival of
D. villosus could irreversibly change these lake
assemblages, even before they have been
documented. To stop or more realistically slow
the invader’s spread, very simple ‘preventative’
measures such as mechanical cleaning and
washing of water sports gear would greatly
reduce the risk factors. For example in the UK, a
recent public education initiative ‘Check, Clean
and Dry’, provides simple guidance for the
public. It recommends simply checking equipment
Dikerogammarus villosus - profile of the invader
and clothing for live organisms. Cleaning and
drying of equipment is then a simple way to
remove/kill invaders and so stop their spread.
F. Efforts to improve river/lake water quality
and wastewater treatment need to be accelerated
as improving water quality would allow native
species to better compete with the invader. This
approach would require considerable economic
resource and political will. One positive aspect
of the D. villosus invasion has been significantly
increased financial support and political/public
interest for research into alien species generally.
G. The killer shrimp is a prime example of an
invader, whose spread has focussed the attention
of the international scientific community and in
particular government agencies tasked with
biodiversity conservation and the protection of
aquatic ecosystems. This interest has been
generated by the relatively early alerts on this
invader’s highly deleterious impacts. Hopefully
forewarned is forearmed.
Acknowledgements
The study was founded by the Polish Ministry for Science and
Higher Education, grant N N304 350139, as well as by internal
grants and funds from the University of Lodz.
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Supplementary material
The following supplementary material is available for this article:
Table S1. Records of Dikerogammarus villosus in native range.
Table S2. Major steps of invasion of Dikerogammarus villosus in Europe.
This material is available as part of online article from:
http://www.aquaticinvasions.net/2014/Supplements/AI_2014_Rewicz_etal_Supplement.xls
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