Introduction
Biological
invasions have become more frequent driven by increasing globalisation,
anthropogenic activities and climate change (Hulme, 2009; Van Kleunen et
al., 2015), inflicting severe negative impacts on local biota, while
imposing economic aftermaths (Gallardo, Clavero, Sánchez, & Vilà, 2016;
Walsh, Carpenter, & Van Der Zanden, 2016). The footprints of biological
invasions can be traced in the changes of species diversity, the
dramatic alterations in communities, habitats, top-down and bottom-up
control modifications, pandemic statuses, shifts in food chains,
nutrient cycling, and in attenuations of ecosystem services (Anton et
al., 2019; David et al., 2017; Simberloff & Rejmánek, 2011). Like in
terrestrial ecosystems, biological invasions in the marine/oceanic realm
are major drivers of ecological and evolutionary shifts, altering
community structures while restructuring ecosystem functions with direct
and indirect impacts on ecosystem services (Carlton & Geller, 1993;
Darling et al., 2017; Katsanevakis et al., 2014). With the increasing
impacts of biological invasions in marine environments, more interest is
being devoted for the inclusion of genetic, phylogenetic and
evolutionary aspects in the research, with parameters that may improve
the resolution and cost effectiveness of monitoring biological invasion
(Darling et al., 2017; Rius, Turon, Bernardi, Volckaert, & Viard, 2015)
and describe changes of population genetics and adaptation properties of
important invasive species in better details (Barrett, 2015; Tepolt,
2015).
Botryllus schlosseri (Pallas, 1766) is a common Mediterranean
Sea/European Atlantic colonial ascidian species, notoriously invasive to
temperate zones worldwide (Reem and Rinkevich, 2014; Freeman,
Frischeisen, & Blakeslee, 2016; Lin and Zhan, 2016; Lord, 2017),
currently inhabiting all continents except Antarctica, including the
coasts of Japan, New Zealand, India, South Africa, Chile, Argentina,
USA, and Canada (Van Name, 1945; Rinkevich, Shapira, Weissman, & Saito,
1992; Ben-Shlomo et al., 2001, 2010). Populations of this species were
assigned to five highly divergent clades on the Cytochrome Oxidase
subunit I (COI) by Bock et al. (2012) with clade A being cosmopolitan,
revealing significant differentiation between native and invasive
populations (Lin & Zhan, 2016), while the other clades are restricted
to European waters. Reem, Douek, Paz, Katzir and Rinkevich (2017)
pointed, however, to the possibility of admixture within two
Mediterranean populations between individuals from different clades,
further revealing that the origin of B. schlosseri , particularly
the cosmopolitan clade A, is still under debate. Early theories (Van
Name, 1945) have suggested that the species originated in European
waters, a proposal supported by Reem et al. (2017), while Yund, Collins
and Johnson (2015) proposed that at least one haplotype in clade A is
native to the northwest Atlantic. Carlton (2005) proposed, yet without
supporting documentations, a possible Pacific origin.
In the United States, B. schlosseri was already present on the
east coast in 1841 (Gould, 1841), while west coast documents (e.g., San
Francisco Bay, Port Hueneme and San Diego, and Bremerton, WA) were
anecdotally recorded around 1944 – 1947 (US Navy, 1951; taxonomic
validation is yet to be performed). In San Francisco area it remained at
low frequencies for at least a decade (Cohen & Carlton, 1995), while
becoming common in San Diego region in the early 1960s (Lambert &
Lambert, 1998). In the Washington area, B. schlosseri was not
reported for the next 36 years, following the first report in 1951 (US
Navy, 1951), until its first validated documentation (Lambert, Lambert
& Kozloff, 1987). Lambert and Lambert (1998) reported an anecdotal
earlier observation of B. schlosseri (late 1960s or early 1970s)
in an oyster farm on San Juan Island, just north of the Puget Sound.
However, subsequent publications did not mention B. schlosseri(Kozloff, 1973, 1974, 1983; Lambert, 1969) in the Washington area. This
species was not recorded by James T. Carlton that surveyed the
biofouling communities in Washington during 1976-1977, and the
mentioning of B. schlosseri by Wonham and Carlton (2005) was a
taxonomic error (J.T. Carlton, personal communication, August 2019).
Furthermore, during 1987, one of the authors of this paper (B.
Rinkevich; unpubl.) performed a survey of marinas along the US west
coast and could not find B. schlosseri north of Coos Bay, Oregon
(an area that was still under active invasion process, as B.
schlosseri failed to establish yet in the upper Coos bay communities;
Hewitt, 1993), nor in several marinas surveyed in the Puget Sound. In
contrast, B. schlosseri was very common in some of the sampled
marinas on his next survey conducted in 1999. The above suggests
possible patchy distributions of early B. schlosseri introduced
to the Seattle area. It is unknown yet whether the current widespread
populations of B. schlosseri in this area (Cohen et al., 1998;
Pleus, Leclair, & Schultz, 2008) originated from earlier invasions,
reported during the 1940’s and 1970’s, or reflect later introductions to
this area.
As an invasive species B. schlosseri negatively impacts
aquaculture organisms (Arens, Christine Paetzold, Ramsay, & Davidson,
2011; Carver, Mallet, & Vercaemer, 2006). Having a short larval
duration of no more than two hours (Grave & Woodbridge, 1924; Grosberg,
1987; Rinkevich & Weissman, 1987), that allows just a limited larval
dispersal (Grosberg, 1987), B. schlosseri relies on floating
objects for long-distance dispersal. It is commonly observed on ship
hulls and aquaculture submersed infrastructure which move between marine
sites (Berrill, 1950; Lambert & Lambert, 1998; Skerman, 1960) and also
spreads to different marinas, probably by boating activity
(Lacoursière-Roussel et al., 2012; López-Legentil, Legentil, Erwin, &
Turon , 2015).
Long-term monitoring of population genetics (García-Navas et al., 2015;
Osborne, Carson, & Turner, 2012) may further elucidate the properties
of an invasive species (Jason Kennington, Hevroy, & Johnson, 2012) and
the roles of their genetic background in dictating the invasiveness
potential of alien species introduced into new territories (Wellband,
Pettitt-Wade, Fisk, & Heath, 2017). Reem, Douek, Katzir and Rinkevich
(2013a) and Karahan, Douek, Paz and Rinkevich (2016) were the first to
study the long-term genetic structures (13 and 12 years, respectively)
of B. schlosseri in Santa Cruz and Moss Landing, two sites in
California, USA, separated by just 20km of shoreline, revealing two
disparate population genetic structures. The Santa Cruz population (Reem
et al., 2013a) is relatively isolated and under high genetic drift,
further characterized by high mutation rates. The Moss Landing
population (Karahan et al., 2016) is affected by episodic freshwater
floods, and subsequent recovery.
These long-term studies on Botryllus schlosseri populations
(Karahan et al., 2016; Reem et al., 2013a) detailed the population
genetic parameters of two populations already established for several
decades. It is therefore of great interest to compare their results with
lately established B. schlosseri populations, like those of the
US north Pacific coast (Puget Sound, Washington), where in the late
1980s this species was missing or just introduced. We studied genetics
parameters of four B. schlosseri populations residing in marinas
in the Seattle WA area, separated by up to a 124 km coastline, during a
period of up to 19 years. Our aim was primarily to get an overall
regional scale perspective of possible genetic structure fluctuations in
recently introduced populations of an invasive species.