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.