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Rapid assessment of non-indigenous species in the era of the eDNA barcoding: A Mediterranean case study
Accepted Manuscript Rapid assessment of non-indigenous species in the era of the eDNA barcoding: A Mediterranean case study Alba Ardura, Serge Planes PII:
S0272-7714(17)30132-4
DOI:
10.1016/j.ecss.2017.02.004
Reference:
YECSS 5380
To appear in:
Estuarine, Coastal and Shelf Science
Received Date: 25 August 2016 Revised Date:
13 January 2017
Accepted Date: 3 February 2017
Please cite this article as: Ardura, A., Planes, S., Rapid assessment of non-indigenous species in the era of the eDNA barcoding: A Mediterranean case study, Estuarine, Coastal and Shelf Science (2017), doi: 10.1016/j.ecss.2017.02.004. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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Rapid assessment of non-indigenous species in the era of the eDNA barcoding: a Mediterranean case study. Alba Ardura* and Serge Planes USR 3278-CRIOBE-CNRS-EPHE, Laboratoire d'excellence “CORAIL”, Université de
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Perpignan-CBETM, 58 Rue Paul Alduy, 66860 Perpignan CEDEX, France *Corresponding author: [email protected]; Tel. +34-660611910 Abstract
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With only a narrow opening through the Gibraltar and Suez Canals, the Mediterranean Sea is one of the largest semi-enclosed seas. The marine flora and fauna are some of the richest in the world, relative to its size, particularly in the coastal habitats, which are
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also characterized by numerous endemic species although the introduction of nonindigenous species threatens its rich and unique biodiversity. Following the opening of the Suez Canal, and in combination with shipping and aquaculture activities, nonindigenous species (NIS) introduction has had measurable impacts on the Mediterranean. Lagoon ecosystems along the French coastline, with approx. 100 NIS identified, are considered hot-spot areas for these species. Rapid assessment sampling
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for sessile benthic species together with DNA barcoding is a rapid, easy and cheap method to detect non-indigenous species. Two nearby and different ecosystems were sampled for invertebrate species: Saint-Nazaire lagoon, a Special Protection Area within the Natura 2000 Network and Canet port, a marina in a small village. The DNA
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barcoding tool for species identification was used for confirming the taxonomy. This showed that, despite the Saint-Nazaire Lagoon classification within the Natura 2000 network, it is already contaminated with a single NIS that was found in high densities
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and is clearly beginning to dominate the system. It is proposed that a rapid assessment of the sampled environment and the DNA barcode approach are efficient and can provide sufficient information on the new target species to be used in conservation planning and ongoing management efforts. Key words: Non-indigenous species (NIS), Mediterranean Sea, Sessile benthic species, DNA barcoding, Rapid assessment, Early detection. Introduction 1
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Estuarine and coastal ecosystems are particularly susceptible to introductions of nonindigenous species (NIS) and many introduced marine and estuarine species have had significant negative effects on communities, ecosystems and resources (Reise et al., 2006; Galil, 2007) and native species (Groshloz, 2002; Williams, 2007; Williams &
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Smith, 2007; Williams & Groshloz, 2008). Coastal and estuarine systems are often heavily impacted by human activities such as shipping and boating (Ruiz et al., 2000; Carlton & Geller, 1993); aquaculture (Naylor et al., 2001); the aquarium trade (Padilla
and Williams, 2004) and live seafood and bait fisheries (Weigel et al., 2005; Chapman et al., 2003), making these areas major vectors for introductions (Williams & Groshloz,
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2008).
The Mediterranean Sea has interacted with human civilizations and their trade routes
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since antiquity, and being a semi-enclosed European sea, exotic species can spread across the entire region very rapidly. A total of 17,000 marine species have been recorded (Coll et al., 2010) in the Mediterranean Sea. This biodiversity accounts for 8 to 9% of the global number of marine species, making the Mediterranean one of the richest seas in the world. In addition, the Mediterranean has a highly diverse coastal zone that supports a high rate of endemism (Fredj et al., 1992). Climate change in conjunction
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with coastal habitat destruction in nearby harbors and lagoons has already driven significant changes in biodiversity. The recent opening of the Suez Canal has led to the introduction of many exotic species, which together with global warming have made the Mediterranean Sea more favorable to tropical species (Goulletquer et al., 2002;
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Katsanevakis et al., 2014; Galil et al., 2015). The introduction of exotic species in the Mediterranean Sea is a dynamic ongoing process with approximately 15 additional species reported each year (EEA Report, 2006). To date, almost 1000 marine non
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indigenous species (NIS) have been introduced in the Mediterranean, of which more than half are considered to be both established and spreading, criteria that categorize them as invasive species. (Zenetos et al., 2010, 2012; Katsanevakis et al., 2014; Galil et al., 2015).
Sessile benthic species are highly adaptable due to their capacity of surviving while submerged and covered. Some of them can be carried by maritime transport very far from the native habitat, spreading to exotic geographic settings where they can become invasive (Smith et al., 2008). The intertidal zone is considered a gateway for the introduction of NIS using intermediate steps to reach new habitats (e.g. Apte et al., 2
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2000). A rapid assessment sampling of this area is a fast and cost-effective method that allows the local detection of NIS (Minchin et al., 2016). Early detection is the best tool to stop new possible invasion (Gozlan et al., 2010; Blanchet, 2012), but it is not always possible, because several species are difficult to detect due to the presence of early
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developmental stages (larvae and juveniles) and cryptic species are also difficult to distinguish from native species. As the accurate identification of species is an essential component of conservation management strategies (Bax et al., 2001) and traditional taxonomic tools do not seem sufficient, DNA barcoding has been cited as the
availability of a reliable, cheap, rapid and accurate tool for NIS identification and
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monitoring (Ardura et al., 2015a; Cross et al., 2010; Briski et al. 2011). DNA-based tools, together with a rapid assessment sampling, allow species identification at any life
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stage, based on DNA extraction from a single individual, facilitating the early detection of new arriving species before an introduced population becomes fully established in a new habitat (Armstrong and Ball, 2005; Chown et al., 2008; Briski et al., 2011; Zhan and MacIsaac, 2015).
Two different locations on the French south Mediterranean coast, Canet port and SaintNazaire lagoon, included in Natura 2000 Network, were sampled randomly. All
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invertebrate specimens obtained were analyzed using a barcoding method that gives a rapid genetic identification of the species (native or NIS), without needing any mature morphological characters, whether vegetative and reproductive. Hence NIS can be detected at any stages of their life cycles, even from morphologically compromised
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samples, such as cryptic species, or highly degraded tissue (Cross et al., 2010). The barcoding was performed with the mitochondrial gene cytochrome oxidase subunit I (COI) which has been intensively used in phylogenetic and phylogeographic works.
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COI sequences have been proven to be very discriminant for in silico genomics (Hebert et al., 2003; Ball et al., 2005). The taxonomic identification was, in some cases, confirmed with a second marker, the 16S rRNA gene. A rapid assessment of one specific target species abundance and distribution within a port has been possible and has provided enough information for decision making in the case of byssate bivalves (Minchin et al., 2016). The aim goal of this study is analyze whether a rapid assessment of biodiversity present in an ecosystem using DNA barcoding tools, can provide sufficient information about the distribution and abundance
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of non-target species (native or non-indigenous species (NIS)) and the ecosystem health. Material and methods
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Study area and sampling Sampling was conducted at 10 sites (Table 1, Fig. 1) in two study areas: 5 sampling sites across the Saint-Nazaire lagoon, a Special Protection Area (SPA) within the Natura
2000 network, and another 5 in Canet port, which is among the largest marinas along
km, with no physical barrier between them.
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the French Mediterranean coast. The two study sites are close to one another, around 4
Canet-Saint Nazaire is a semi-closed system and because of this, like many coastal
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lagoons spanning the French Mediterranean coastline, it is highly susceptible to NIS introduction. Interfacing between catchments and the sea, these transitional aquatic ecosystems are exposed to important seasonal variations of temperature and to sudden and intense Mediterranean rainfalls that lead to flash floods bringing large volumes of freshwater into lagoons, thus reducing their salinity (Pecqueur et al., 2011). This particular lagoon is part of the greater Canet-Saint Nazaire lagoon network with an
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average salinity of 20.4 (Bec et al., 2011) and temperature of 13.5ºC (Global Sea Temperature website). The habitat diversity, from grasslands to sandy dunes, grazed through wetlands and a varied salinity gradient, contribute to the area’s high biodiversity. It is located along one of Europe’s principal migration routes and since March 2006 it has been classified as Special Protection Area (SPA). However, due to
2010).
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urban eutrophication, it is much degraded with a hypertrophic status (Souchu et al.,
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Within 4 km north of the lagoon complex, there is a busy harbor: Canet port. It is a very popular location for boating related tourism; Canet has space for 977 boats up to 24 m in length (Portbooker website). In this area there are more berths per mile that anywhere else in the French Mediterranean. The salinity is higher than in Saint Nazaire lagoon, with 37.5 average salinity (Bec et al., 2011) and little warmer with an average temperature of 15.9ºC (Global Sea Temperature website). It has no eutrophication problems but it shows urban pollution from the nearby small village of Canet (Souchu et al., 2010).
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The sampling protocol was based on the Rapid Assessment (RAS) approach described by Minchin (2007). Sampling involved the collection of all invertebrate specimens that could be distinguished de visu with a maximum of 100 samples per study area (Table 2). In Canet port, the samples taken were attached on the rocky area or artificial
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structures on the surface and at 2 m depth accessible by Scuba diving. In the SaintNazaire lagoon, samples were taken from the sand using a sieve. For representative
sampling, and preventing the biased collection of species with a patchy distribution, a
visual inspection prior to sampling was made to determine that phenotypically different
organisms (presumably different species) were present in the sampling site. The number
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of individuals picked of each morphotype was approximately proportional to the
abundance of the respective morphotype. To standardize the sampling effort, each
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location was sampled for 1 hour in total for each of the 5 sites or until 100 individuals were collected.
All individuals were immediately transported in ambient water (in 1 L buckets) to the Centre de Biologie et d’Ecologie Tropicale et Méditerranéenne (University of Perpignan) for visual species identification and subsequent preservation for molecular analyses. In the laboratory, the specimens sampled were visually identified employing
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taxonomic guides. Next, a piece of tissue from each individual was preserved in absolute ethanol for further genetic analysis. One specimen from each species was stored in ethanol as a voucher specimen.
Classification of species as Non indigenous Species (NIS)
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The species found in this study were classified either as native or NIS, according to the native distribution of each species (WoRMS, 2016). The Invasive Species Specialist
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Group (ISSG) global database (ISSG, 2016) of the International Union for Conservation of Nature (IUCN) was used as the reference to check the invasiveness status and invasion histories of each identified non-indigenous species. The taxonomic nomenclature of all identified species was verified against the World Register of Marine Species (WoRMS, 2016). DNA extractions and PCR designs Total DNA was extracted from a small piece of tissue with the E.Z.N.A Mollusk DNA kit (IOMEGA, bio-tek) in order to remove high contents of mucopolysaccharides in muscle tissues, following manufacturer instructions. For the other invertebrates, total 5
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DNA was extracted following the standard protocol described by Estoup et al. (1996), employing Chelex® resin (Bio-Rad Laboratories). DNA was stored at 4ºC for immediate DNA analysis, and aliquots were frozen at -20ºC for long-term preservation. A fragment of about 600 bp of the COI sequence was amplified by polymerase chain
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reaction (PCR), employing the primers described by Geller et al. (2013). The amplification reaction was performed in a total volume of 23 µl, with 5 PRIME Buffer,
1x (Gaithersburg, MD, USA), 1.5 mM MgCl2, 0.25 mM dNTPs, 1 µM of each primer,
approximately 20 ng of template DNA and 1.5 U of DNA Taq polymerase (5 PRIME),
and the following PCR conditions: initial denaturing at 95ºC for 5 minutes, 35 cycles of
minute and final extension at 72ºC for 5 minutes.
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denaturing at 95ºC for 1 minute, annealing at 48ºC for 1 minute, extension at 72ºC for 1
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Additional sequencing of about 550 bp of the mitochondrial 16S rRNA sequence with the primers described by Palumbi (1996) was carried out for some samples in order to complete the taxonomic identification with a second marker. The amplification reaction was performed in a total volume of 23 µl with the same conditions described above for the COI gene and the following PCR conditions: initial denaturing at 95ºC for 5 minutes, 30 cycles of denaturing at 94ºC for 1 minute, annealing at 55ºC for 1 minute,
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extension at 72ºC for 2 minutes and a final extension at 72ºC for 7 minutes. PCR products were visualized in 2% agarose gels with 3 µl of 10 mg/ml ethidium bromide. Sequencing was performed externally with the DNA sequencing service Genoscreen (France).
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Sequence edition
Sequences were visualized and edited employing the BioEdit Sequence Alignment
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Editor software (Hall, 1999) and aligned with the ClustalW application (Thompson et al., 1994) included in BioEdit. Sequences obtained from our screening were compared with international databases employing BLAST within NCBI (NCBI, 2016) and BOLD system, in the case of COI sequences (BOLD, 2016), for species identification. Diversity indices
Species diversity itself has two separate components: the number of species present (species richness) and their relative abundance (termed dominance or evenness). Different diversity indices for species richness (total number of species and Margalef index), their relative abundance (Simpson index) and phylogenetic structure 6
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(phylogenetic diversity) were calculated using PRIMER 6 software, which together provide the baseline information regarding the biodiversity status for a specific area: -
Total number of species (NSp), the number of species in each sample. i.e.
species with non-zero counts. Species richness using Margalef index (SR (d)). d = (S-1)/Log (N) - Margalef's
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-
species richness for each sample is a measure of the number of species present, making some allowance for the number of individuals. -
Ecological biodiversity using Simpson index (1-λ´= 1-SUM (Ni*(Ni-1)/(N*(N-
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1))). Since evenness and dominance are complimentary, this index takes into account the number of species in the habitat and their relative abundance; it is based on the
probability of any two individuals drawn at random from an infinitely large community
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belonging to the same species. Simpson index is usually reported as its complement (1λ´), because λ´ is a measure of dominance, so as λ´ increases, diversity (in the sense of evenness) decreases. -
Total phylogenetic diversity to determine the phylogenetic structure at each
sampling point (PD). Phylogenetic diversity is a measure of taxonomic distinctiveness;
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it is the total path length constituting the full taxonomic tree.
For all of the indices, higher values represent higher biodiversity in sense of species richness, relative abundance or phylogenetic structure. Results
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In total, 200 macrofauna specimens were sequenced from 5 sites in each of the two locations. Overall, DNA barcodes from 28 invertebrate species were obtained
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confirming the morphological identification at the species level in 25 cases; with 152 DNA sequences from the COI gene of about 600 nucleotides long and 48 sequences from 16S rRNA genes of 550 pb for some taxa. One haplotype per species was submitted to GenBank database, and they are now available with accession numbers in the NCBI database (Table 1). From BLAST analysis, these sequences were unequivocally assigned to 25 different species (Table 2) with E-values <<0.0001 in all cases. Both genes confirmed the assignation of 25 species. The diversity indices (NSp; SR (d), Simpson and PD) varied between the two sampling locations (Table 3), and in all cases were higher in Canet port than in Saint-Nazaire 7
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Lagoon. Four species were identified as NIS (Table 2) that accounted for 22% of the 200 barcoded individuals. Three NIS correspond to Canet port sites while only one was collected from the Saint-Nazaire Lagoon; corresponding to 12.5% and 20% of the total diversity respectively of the total invertebrate community.
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Three of the NIS identified, Crassostrea gigas, Ficopomatus enigmaticus and Styela
plicata, have wide distributions and their effects on ecosystems have been previously reported in the literature. The last species, Aiptasia pulchella, is more restricted in its
distribution, found only in Atlantic Europe and the French Mediterranean basin. The most abundant NIS was F. enigmaticus (a tubeworm probably native to Australia,
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(McQuaid and Griffiths, 2014)) forming 35% of the samples analyzed in Saint-Nazaire Lagoon. The other three NIS were present in much smaller proportions in Canet port:
pulchella (brown sea anemone). Discussion
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4% of C. gigas (Pacific oyster), 4% of S. plicata (pleated sea squirt) and 1% of A.
The rapid assessment method based on estimation of the abundance and distribution of target species may be conducted in a short time in a very cheap, fast an easy way (Minchin et al., 2016). However, we cannot always know the target species, especially
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with unknown new arriving species that can be cryptic species that are very similar to native ones. This problem is very important in ballast water management procedures, for which the continuous incorporation of data from new or updated biological surveys is essential to develop a good target species database in order to have a scientifically
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robust risk assessment (Olenin et al., 2016). For this purpose, DNA barcoding is essential to describe newly arriving species, because we cannot identify an organism
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looking at tissues and in some cases, therefore a traditional taxonomic approach is not sufficient to classify new species. In addition, the time and money needed for a morphological approach is much greater than that necessary for DNA barcoding. Only a small amount of tissue is required, even degraded (the whole organism is not necessary), in any develop stage (eggs, larvae, adults...). In this work a rapid random assessment sampling together with DNA barcoding tool have been used and emphasizes the value of DNA barcoding as a tool for biodiversity inventory (Ardura et al., 2011). This method is especially applicable to marine invertebrates, as there are many reference sequences available from many different geographical areas (Hebert et al., 2003; Bucklin et al., 2011; Geller et al., 2013). However, in some cases the database is 8
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not sufficient for assigning species as occurs with the three phenotypes from our sampling: they can be assigned at genus level (Table 2); in these cases a previous taxonomic work or sequencing of other genera is needed, but it is not the aim of this study, especially since they are not NIS and the genus is sufficient for biodiversity
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analysis. In this particular work, genetic barcoding helps to solve uncertainties, especially in ascidian species in which microscopic examination is usually necessary for species identification (Hirose & Hirose, 2013).
Both ecosystems sampled in this study have different natural and anthropogenic
characteristics. The Saint-Nazaire lagoon is less saline and has a lower temperature on
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average and it is involved in the Natura 2000 Network, however it presents serious eutrophication problems due to human activity (Souchu et al., 2010). On the other hand,
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Canet port sustains intense recreational boat traffic, without eutrophication problems but with urban pollution (Souchu et al., 2010). The diversity calculated based on different parameters (species richness, dominance and phylogenetic structure) is always greater in Canet port than in Saint-Nazaire lagoon (Table 3) and can be used as proxy biodiversity indicator (May, 1995; Bianchi & Morri, 2000; Katsanevakis et al. 2014). Three NIS species were found in Canet port while only one was found in Saint-Nazaire
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lagoon (Table 3). The presence of a greater number of NIS in Canet port is what would be expected of a busy harbor with significant maritime traffic (Hewitt & Martin, 1996; Firestone and Corbett, 2005; Leuven et al., 2009). On the other hand, the presence of Ficopomatus enigmaticus within the Saint-Nazaire Special Protection Area questions
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the efficacy of protected areas as a means of preventing the introduction of NIS. This introduced species was probably transported on the hulls of ships or on commercial mollusk shells (de Wit, 2011). It is most prominent and grows best in estuaries and
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lagoons with brackish waters and a high nutrient content (JNCC, 1997; Schwind & Iribarne, 2000; Bianchi & Morri, 2001; Paavola et al., 2005). The three NIS species present in Canet port, Crassostrea gigas, Aiptasia pulchella and Styela plicata, are native from the NW Pacific, probably originating from Japanese coasts. These species have already been described along French coasts (Gruet et al., 1976), and their introductions are most likely a result of the ‘controlled’ authorized introduction of C. gigas for aquaculture purposes from Japanese stocks (2001). The NIS identified in Saint-Nazaire Lagoon, Ficopomatus enigmaticus, is presumably native to Australia (McQuaid and Griffiths, 2014). It was first noticed in northern France in 1921 (de Wit, 9
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2011) and a previous study by Fornos et al. (1997) described its presence in the Western Mediterranean lagoons as well. It is probable that it was introduced either through transport on the hulls of ships or on commercial mollusk shells (de Wit, 2011). The reef-building polychaete F. enigmaticus has been shown to have major impacts on the
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systems it invades and is therefore considered an ecosystem engineer (Schwindt et al., 2001, 2004). Their impacts have been shown worldwide (i.e. Mar Chiquita coastal lagoon in Argentina (Schwindt et al., 2004) or in Zandvlei Estuary, Cape Town, South Africa (McQuaid and Griffiths, 2014)). This polychaete clearly impacts
native
biodiversity, its presence in the Saint-Nazaire lagoon represents to 20% of the total of
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biodiversity inventoried in this area (Table 3) this proportion being a signal of
significant biodiversity turnover (Katsanevakis et al., 2014), and it is higher than that
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found in Canet port (12.5%) corresponding to three different NIS.
Protected areas are a common tool for the conservation of biodiversity, vulnerable species, ecological processes and the protection of resources from overexploitation and from the destructive effects of human activities in both terrestrial and marine environments (Gaines et al., 2010; Pimm et al., 2001; Le Saout et al., 2013; Rife et al., 2013). The Natura 2000 network is currently the centerpiece of EU nature &
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biodiversity conservation policy, but as the data demonstrate that the mere establishment of Protected Areas does not guarantee their success. When protected areas are simply decreed but insufficient resources are available for effective design, management, or enforcement, these become only “paper parks” that do not effectively
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restrict exploitation or access (White & Courtney, 2004; Rife et al., 2013). For the Canet Saint-Nazaire Lagoon, this is the case, there is currently no management plan for the lagoon (see site code: FR9112025, Natura2000, 2016).
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The biggest concern is that estuaries have been urbanized worldwide as the popularity of coastal areas increases continuously (Chapman, 2006; Bockley, 2007) and their environmental protection is affected by economic interests. Natural ecosystems are being replaced by man-made ones with hard surfaces replacing soft sand or sediments (Bulleri and Airoldi, 2005; Chapman, 2006; Blockley, 2007; Glasby et al., 2007; Dafforn et al., 2009). These artificial structures are an ideal substrate for new coming species, normally dominated by NIS (Bulleri and Airoldi, 2005; Glasby et al., 2007). The newly arriving species, together with new artificial structures, are altering the habitat by increasing habitat complexity with an influence on the abundance, diversity 10
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and distribution of the organism chain (Coull and Wells, 1983) and native biodiversity is being threatened with habitat destruction and the introduction of NIS (Glasby et al., 2007). Therefore, habitat degradation is more of a promoter than a consequence of biological invasions (MacDougall and Turkington, 2005), and similar to terrestrial
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ecosystems, species richness of marine communities is negatively correlated with invasion success (Stachowicz et al., 1999). Thus, native biodiversity and habitat health
could be strong barriers for biological invasions. While tourism is a very important economic activity for many areas worldwide, some interventions can be carried out to
avoid the arrival and dispersion of new coming species, such as cleaning recreational
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vessels before transport between estuaries, avoiding the dumping of hard objects to limit the availability of suitable substrata for F. enigmaticus and other similar non-indigenous
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invertebrate species (McQuaid and Griffiths, 2004).
The methodology developed in this study is very valuable for a rapid assessment of invertebrate species biodiversity, because the individuals can be taken easily from the environment and it is a very fast, cheap and easy technique (without the need for very specialized human and technical resources) to inventory biodiversity and detect NIS. The problem is that sampling is still necessary and requires a large effort of human
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resources and many specialists systematically sampling all ecosystems, which is even greater in habitats with difficult physical access that have to be accessed from the sea and/or diving. In addition, some species cannot be detected because they are at a low density, in their first development stages or with higher mobility (not sessile species). In
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these cases, the use of environmental DNA, extracted directly from water samples, together with high-throughput sequencing (HTS) metabarcoding is capable of nontarget species detection and is highly sensitive (Ardura et al., 2016; Ardura et al.,
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2015b; Zaiko et al. 2015). However, the cost of the sequencing is significantly higher and substantial analytical effort (bioinformatics) is needed to ensure efficient exploration of the sequence data obtained from multi-species communities (Blanchet, 2012). For this approach taxonomic expertise is not required, eDNA-HTS techniques can supplement observational records and field surveys to obtain marine ecosystem samples and information, as they reveal the number of NIS in one specific area and their temporal and spatial distribution (Ardura et al., 2015; Zaiko et al., 2015). The best methodology should be assessed for each particular study, depending on economic and material resources and the data available and necessary in each case.
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Given the above, with one or another methodology, the early detection of NIS (target or not-target species), its knowledge about their distribution and abundance and the biodiversity inventory are essential to know the state and health of a particular environment. This knowledge is necessary for the management of NIS and it should go
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together with efforts and resources that nations have applied to reduce pollution and to restore wetlands and fishery stocks, as all of these elements are interconnected and they are promoters and consequences of the same problem: the anthropogenic impact on the
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environment.
Acknowledgments
Thank you to Martin Desmalades for helping with the sampling tasks and to Jeanine
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Almany for improvements on the manuscript. We are grateful to Thomas E. Holloway (Royal Holloway University of London, UK) for proofreading the manuscript. A.A. holds a regional postdoctoral Marie Curie grants COFUND-CLARIN. References
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Fig. 1. Map of the sampling points. Saint-Nazaire Lagoon and Canet port situated on the Mediterranean Sea, and the 5 points sampled of each location.
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Table 1. Sampling sites with GPS coordinates.
Latitude
Longitude
Saint-Nazaire-1
A1
42.67413° N
3.0134º E
Saint-Nazaire-2
A2
42.67641º N
3.00027º E
Saint-Nazaire-3
A3
42.67204º N
3.03250º E
Saint-nazaire-4
A4
42.65949º N
3.03439º E
Saint-Nazaire-5
A5
42.65261º N
Canet port-1
B1
42.70201º N
Canet port-3 Canet port-4
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B2
42.70433º N
3.03516º E
B3
42.70468º N
3.02776º E
B4
42.70414° N
3.03284° E
B5
42.70360° N
3.03845° E
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Canet port-5
3.03158º E
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Canet port-2
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Point
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Table 2. Species found from each sampling location, in percent (N= sample size). Nonindigenous species are marked in bold. GenBank Accesion Number (GB-AN).
Saint-Nazaire
Abra sp
Native
0.37
Aiptasia pulchella
NW Pacific
0.01
Anomia ephippium
Native
0.04
Balanus perforatus
Native
0.05
Carcinus aestuarii
Native
0.01
Cerastoderma glaucum
Native
Crassostrea gigas
NW Pacific
Ficopomatus enigmaticus
Australia
Gibbula adansonii
Native
Gibbula varia
Native
Hediste diversicolor
Native
Lumbrineris funchalensis
Native
Mimachlamys varia
16S GB-AN
-
KT988336
KT988316
-
KT988315
-
KT970486
-
-
KT988319
-
KT988318
-
-
KT988337
0.01
KT988320
-
0.05
KT988321
-
KT988322
-
0.01
-
KT988339
Native
0.01
KT988323
-
Native
0.05
-
KT988340
Native
0.29
KT988324
-
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Mytilus galloprovincialis
COI GB-AN
KT988317
0.01
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0.04
0.35
0.26
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Maja crispate
Canet port
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Status
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Species
Native
0.01
KT988325
-
Nassarius sp
Native
0.03
-
KT988341
Ocenebra erinaceus
Native
0.01
KT988327
-
Osilinus turbinatus
Native
0.04
KT988328
-
Ostrea edulis
Native
0.04
KT988326
-
Pachygrapsus marmoratus
Native
0.01
KT988331
-
Pagurus cuanensis
Native
0.01
KT988329
-
Patella caerulea
Native
0.18
KT988330
-
Patella rustica
Native
0.02
KT988332
-
Patella ulyssiponensis
Native
0.03
KT988333
-
Polynoe sp
Native
0.01
-
KT988342
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Nassarius incrassatus
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Sabella pavonina
Native
0.01
KT988334
-
Styela plicata
NW Pacific
0.04
KT988335
-
100
100
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N
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Table 3. Diversity indices values for invertebrate communities found in the two considered locations on the Mediterranean Sea. Number of species (NSp), species richness (SR), proportion of NIS (NIS), Simpson’s diversity index and phylogenetic diversity (PD).
SR (d)
NIS
Simpson
PD
Saint-Nazaire
5
0.869
0.200
0.680
416.7
Canet port
24
4.994
0.125
0.882
1,433
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NSp
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S0272-7714(17)30132-4
DOI:
10.1016/j.ecss.2017.02.004
Reference:
YECSS 5380
To appear in:
Estuarine, Coastal and Shelf Science
Received Date: 25 August 2016 Revised Date:
13 January 2017
Accepted Date: 3 February 2017
Please cite this article as: Ardura, A., Planes, S., Rapid assessment of non-indigenous species in the era of the eDNA barcoding: A Mediterranean case study, Estuarine, Coastal and Shelf Science (2017), doi: 10.1016/j.ecss.2017.02.004. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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Rapid assessment of non-indigenous species in the era of the eDNA barcoding: a Mediterranean case study. Alba Ardura* and Serge Planes USR 3278-CRIOBE-CNRS-EPHE, Laboratoire d'excellence “CORAIL”, Université de
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Perpignan-CBETM, 58 Rue Paul Alduy, 66860 Perpignan CEDEX, France *Corresponding author: [email protected]; Tel. +34-660611910 Abstract
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With only a narrow opening through the Gibraltar and Suez Canals, the Mediterranean Sea is one of the largest semi-enclosed seas. The marine flora and fauna are some of the richest in the world, relative to its size, particularly in the coastal habitats, which are
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also characterized by numerous endemic species although the introduction of nonindigenous species threatens its rich and unique biodiversity. Following the opening of the Suez Canal, and in combination with shipping and aquaculture activities, nonindigenous species (NIS) introduction has had measurable impacts on the Mediterranean. Lagoon ecosystems along the French coastline, with approx. 100 NIS identified, are considered hot-spot areas for these species. Rapid assessment sampling
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for sessile benthic species together with DNA barcoding is a rapid, easy and cheap method to detect non-indigenous species. Two nearby and different ecosystems were sampled for invertebrate species: Saint-Nazaire lagoon, a Special Protection Area within the Natura 2000 Network and Canet port, a marina in a small village. The DNA
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barcoding tool for species identification was used for confirming the taxonomy. This showed that, despite the Saint-Nazaire Lagoon classification within the Natura 2000 network, it is already contaminated with a single NIS that was found in high densities
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and is clearly beginning to dominate the system. It is proposed that a rapid assessment of the sampled environment and the DNA barcode approach are efficient and can provide sufficient information on the new target species to be used in conservation planning and ongoing management efforts. Key words: Non-indigenous species (NIS), Mediterranean Sea, Sessile benthic species, DNA barcoding, Rapid assessment, Early detection. Introduction 1
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Estuarine and coastal ecosystems are particularly susceptible to introductions of nonindigenous species (NIS) and many introduced marine and estuarine species have had significant negative effects on communities, ecosystems and resources (Reise et al., 2006; Galil, 2007) and native species (Groshloz, 2002; Williams, 2007; Williams &
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Smith, 2007; Williams & Groshloz, 2008). Coastal and estuarine systems are often heavily impacted by human activities such as shipping and boating (Ruiz et al., 2000; Carlton & Geller, 1993); aquaculture (Naylor et al., 2001); the aquarium trade (Padilla
and Williams, 2004) and live seafood and bait fisheries (Weigel et al., 2005; Chapman et al., 2003), making these areas major vectors for introductions (Williams & Groshloz,
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2008).
The Mediterranean Sea has interacted with human civilizations and their trade routes
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since antiquity, and being a semi-enclosed European sea, exotic species can spread across the entire region very rapidly. A total of 17,000 marine species have been recorded (Coll et al., 2010) in the Mediterranean Sea. This biodiversity accounts for 8 to 9% of the global number of marine species, making the Mediterranean one of the richest seas in the world. In addition, the Mediterranean has a highly diverse coastal zone that supports a high rate of endemism (Fredj et al., 1992). Climate change in conjunction
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with coastal habitat destruction in nearby harbors and lagoons has already driven significant changes in biodiversity. The recent opening of the Suez Canal has led to the introduction of many exotic species, which together with global warming have made the Mediterranean Sea more favorable to tropical species (Goulletquer et al., 2002;
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Katsanevakis et al., 2014; Galil et al., 2015). The introduction of exotic species in the Mediterranean Sea is a dynamic ongoing process with approximately 15 additional species reported each year (EEA Report, 2006). To date, almost 1000 marine non
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indigenous species (NIS) have been introduced in the Mediterranean, of which more than half are considered to be both established and spreading, criteria that categorize them as invasive species. (Zenetos et al., 2010, 2012; Katsanevakis et al., 2014; Galil et al., 2015).
Sessile benthic species are highly adaptable due to their capacity of surviving while submerged and covered. Some of them can be carried by maritime transport very far from the native habitat, spreading to exotic geographic settings where they can become invasive (Smith et al., 2008). The intertidal zone is considered a gateway for the introduction of NIS using intermediate steps to reach new habitats (e.g. Apte et al., 2
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2000). A rapid assessment sampling of this area is a fast and cost-effective method that allows the local detection of NIS (Minchin et al., 2016). Early detection is the best tool to stop new possible invasion (Gozlan et al., 2010; Blanchet, 2012), but it is not always possible, because several species are difficult to detect due to the presence of early
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developmental stages (larvae and juveniles) and cryptic species are also difficult to distinguish from native species. As the accurate identification of species is an essential component of conservation management strategies (Bax et al., 2001) and traditional taxonomic tools do not seem sufficient, DNA barcoding has been cited as the
availability of a reliable, cheap, rapid and accurate tool for NIS identification and
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monitoring (Ardura et al., 2015a; Cross et al., 2010; Briski et al. 2011). DNA-based tools, together with a rapid assessment sampling, allow species identification at any life
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stage, based on DNA extraction from a single individual, facilitating the early detection of new arriving species before an introduced population becomes fully established in a new habitat (Armstrong and Ball, 2005; Chown et al., 2008; Briski et al., 2011; Zhan and MacIsaac, 2015).
Two different locations on the French south Mediterranean coast, Canet port and SaintNazaire lagoon, included in Natura 2000 Network, were sampled randomly. All
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invertebrate specimens obtained were analyzed using a barcoding method that gives a rapid genetic identification of the species (native or NIS), without needing any mature morphological characters, whether vegetative and reproductive. Hence NIS can be detected at any stages of their life cycles, even from morphologically compromised
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samples, such as cryptic species, or highly degraded tissue (Cross et al., 2010). The barcoding was performed with the mitochondrial gene cytochrome oxidase subunit I (COI) which has been intensively used in phylogenetic and phylogeographic works.
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COI sequences have been proven to be very discriminant for in silico genomics (Hebert et al., 2003; Ball et al., 2005). The taxonomic identification was, in some cases, confirmed with a second marker, the 16S rRNA gene. A rapid assessment of one specific target species abundance and distribution within a port has been possible and has provided enough information for decision making in the case of byssate bivalves (Minchin et al., 2016). The aim goal of this study is analyze whether a rapid assessment of biodiversity present in an ecosystem using DNA barcoding tools, can provide sufficient information about the distribution and abundance
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of non-target species (native or non-indigenous species (NIS)) and the ecosystem health. Material and methods
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Study area and sampling Sampling was conducted at 10 sites (Table 1, Fig. 1) in two study areas: 5 sampling sites across the Saint-Nazaire lagoon, a Special Protection Area (SPA) within the Natura
2000 network, and another 5 in Canet port, which is among the largest marinas along
km, with no physical barrier between them.
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the French Mediterranean coast. The two study sites are close to one another, around 4
Canet-Saint Nazaire is a semi-closed system and because of this, like many coastal
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lagoons spanning the French Mediterranean coastline, it is highly susceptible to NIS introduction. Interfacing between catchments and the sea, these transitional aquatic ecosystems are exposed to important seasonal variations of temperature and to sudden and intense Mediterranean rainfalls that lead to flash floods bringing large volumes of freshwater into lagoons, thus reducing their salinity (Pecqueur et al., 2011). This particular lagoon is part of the greater Canet-Saint Nazaire lagoon network with an
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average salinity of 20.4 (Bec et al., 2011) and temperature of 13.5ºC (Global Sea Temperature website). The habitat diversity, from grasslands to sandy dunes, grazed through wetlands and a varied salinity gradient, contribute to the area’s high biodiversity. It is located along one of Europe’s principal migration routes and since March 2006 it has been classified as Special Protection Area (SPA). However, due to
2010).
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urban eutrophication, it is much degraded with a hypertrophic status (Souchu et al.,
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Within 4 km north of the lagoon complex, there is a busy harbor: Canet port. It is a very popular location for boating related tourism; Canet has space for 977 boats up to 24 m in length (Portbooker website). In this area there are more berths per mile that anywhere else in the French Mediterranean. The salinity is higher than in Saint Nazaire lagoon, with 37.5 average salinity (Bec et al., 2011) and little warmer with an average temperature of 15.9ºC (Global Sea Temperature website). It has no eutrophication problems but it shows urban pollution from the nearby small village of Canet (Souchu et al., 2010).
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The sampling protocol was based on the Rapid Assessment (RAS) approach described by Minchin (2007). Sampling involved the collection of all invertebrate specimens that could be distinguished de visu with a maximum of 100 samples per study area (Table 2). In Canet port, the samples taken were attached on the rocky area or artificial
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structures on the surface and at 2 m depth accessible by Scuba diving. In the SaintNazaire lagoon, samples were taken from the sand using a sieve. For representative
sampling, and preventing the biased collection of species with a patchy distribution, a
visual inspection prior to sampling was made to determine that phenotypically different
organisms (presumably different species) were present in the sampling site. The number
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of individuals picked of each morphotype was approximately proportional to the
abundance of the respective morphotype. To standardize the sampling effort, each
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location was sampled for 1 hour in total for each of the 5 sites or until 100 individuals were collected.
All individuals were immediately transported in ambient water (in 1 L buckets) to the Centre de Biologie et d’Ecologie Tropicale et Méditerranéenne (University of Perpignan) for visual species identification and subsequent preservation for molecular analyses. In the laboratory, the specimens sampled were visually identified employing
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taxonomic guides. Next, a piece of tissue from each individual was preserved in absolute ethanol for further genetic analysis. One specimen from each species was stored in ethanol as a voucher specimen.
Classification of species as Non indigenous Species (NIS)
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The species found in this study were classified either as native or NIS, according to the native distribution of each species (WoRMS, 2016). The Invasive Species Specialist
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Group (ISSG) global database (ISSG, 2016) of the International Union for Conservation of Nature (IUCN) was used as the reference to check the invasiveness status and invasion histories of each identified non-indigenous species. The taxonomic nomenclature of all identified species was verified against the World Register of Marine Species (WoRMS, 2016). DNA extractions and PCR designs Total DNA was extracted from a small piece of tissue with the E.Z.N.A Mollusk DNA kit (IOMEGA, bio-tek) in order to remove high contents of mucopolysaccharides in muscle tissues, following manufacturer instructions. For the other invertebrates, total 5
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DNA was extracted following the standard protocol described by Estoup et al. (1996), employing Chelex® resin (Bio-Rad Laboratories). DNA was stored at 4ºC for immediate DNA analysis, and aliquots were frozen at -20ºC for long-term preservation. A fragment of about 600 bp of the COI sequence was amplified by polymerase chain
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reaction (PCR), employing the primers described by Geller et al. (2013). The amplification reaction was performed in a total volume of 23 µl, with 5 PRIME Buffer,
1x (Gaithersburg, MD, USA), 1.5 mM MgCl2, 0.25 mM dNTPs, 1 µM of each primer,
approximately 20 ng of template DNA and 1.5 U of DNA Taq polymerase (5 PRIME),
and the following PCR conditions: initial denaturing at 95ºC for 5 minutes, 35 cycles of
minute and final extension at 72ºC for 5 minutes.
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denaturing at 95ºC for 1 minute, annealing at 48ºC for 1 minute, extension at 72ºC for 1
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Additional sequencing of about 550 bp of the mitochondrial 16S rRNA sequence with the primers described by Palumbi (1996) was carried out for some samples in order to complete the taxonomic identification with a second marker. The amplification reaction was performed in a total volume of 23 µl with the same conditions described above for the COI gene and the following PCR conditions: initial denaturing at 95ºC for 5 minutes, 30 cycles of denaturing at 94ºC for 1 minute, annealing at 55ºC for 1 minute,
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extension at 72ºC for 2 minutes and a final extension at 72ºC for 7 minutes. PCR products were visualized in 2% agarose gels with 3 µl of 10 mg/ml ethidium bromide. Sequencing was performed externally with the DNA sequencing service Genoscreen (France).
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Sequence edition
Sequences were visualized and edited employing the BioEdit Sequence Alignment
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Editor software (Hall, 1999) and aligned with the ClustalW application (Thompson et al., 1994) included in BioEdit. Sequences obtained from our screening were compared with international databases employing BLAST within NCBI (NCBI, 2016) and BOLD system, in the case of COI sequences (BOLD, 2016), for species identification. Diversity indices
Species diversity itself has two separate components: the number of species present (species richness) and their relative abundance (termed dominance or evenness). Different diversity indices for species richness (total number of species and Margalef index), their relative abundance (Simpson index) and phylogenetic structure 6
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(phylogenetic diversity) were calculated using PRIMER 6 software, which together provide the baseline information regarding the biodiversity status for a specific area: -
Total number of species (NSp), the number of species in each sample. i.e.
species with non-zero counts. Species richness using Margalef index (SR (d)). d = (S-1)/Log (N) - Margalef's
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-
species richness for each sample is a measure of the number of species present, making some allowance for the number of individuals. -
Ecological biodiversity using Simpson index (1-λ´= 1-SUM (Ni*(Ni-1)/(N*(N-
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1))). Since evenness and dominance are complimentary, this index takes into account the number of species in the habitat and their relative abundance; it is based on the
probability of any two individuals drawn at random from an infinitely large community
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belonging to the same species. Simpson index is usually reported as its complement (1λ´), because λ´ is a measure of dominance, so as λ´ increases, diversity (in the sense of evenness) decreases. -
Total phylogenetic diversity to determine the phylogenetic structure at each
sampling point (PD). Phylogenetic diversity is a measure of taxonomic distinctiveness;
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it is the total path length constituting the full taxonomic tree.
For all of the indices, higher values represent higher biodiversity in sense of species richness, relative abundance or phylogenetic structure. Results
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In total, 200 macrofauna specimens were sequenced from 5 sites in each of the two locations. Overall, DNA barcodes from 28 invertebrate species were obtained
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confirming the morphological identification at the species level in 25 cases; with 152 DNA sequences from the COI gene of about 600 nucleotides long and 48 sequences from 16S rRNA genes of 550 pb for some taxa. One haplotype per species was submitted to GenBank database, and they are now available with accession numbers in the NCBI database (Table 1). From BLAST analysis, these sequences were unequivocally assigned to 25 different species (Table 2) with E-values <<0.0001 in all cases. Both genes confirmed the assignation of 25 species. The diversity indices (NSp; SR (d), Simpson and PD) varied between the two sampling locations (Table 3), and in all cases were higher in Canet port than in Saint-Nazaire 7
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Lagoon. Four species were identified as NIS (Table 2) that accounted for 22% of the 200 barcoded individuals. Three NIS correspond to Canet port sites while only one was collected from the Saint-Nazaire Lagoon; corresponding to 12.5% and 20% of the total diversity respectively of the total invertebrate community.
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Three of the NIS identified, Crassostrea gigas, Ficopomatus enigmaticus and Styela
plicata, have wide distributions and their effects on ecosystems have been previously reported in the literature. The last species, Aiptasia pulchella, is more restricted in its
distribution, found only in Atlantic Europe and the French Mediterranean basin. The most abundant NIS was F. enigmaticus (a tubeworm probably native to Australia,
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(McQuaid and Griffiths, 2014)) forming 35% of the samples analyzed in Saint-Nazaire Lagoon. The other three NIS were present in much smaller proportions in Canet port:
pulchella (brown sea anemone). Discussion
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4% of C. gigas (Pacific oyster), 4% of S. plicata (pleated sea squirt) and 1% of A.
The rapid assessment method based on estimation of the abundance and distribution of target species may be conducted in a short time in a very cheap, fast an easy way (Minchin et al., 2016). However, we cannot always know the target species, especially
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with unknown new arriving species that can be cryptic species that are very similar to native ones. This problem is very important in ballast water management procedures, for which the continuous incorporation of data from new or updated biological surveys is essential to develop a good target species database in order to have a scientifically
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robust risk assessment (Olenin et al., 2016). For this purpose, DNA barcoding is essential to describe newly arriving species, because we cannot identify an organism
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looking at tissues and in some cases, therefore a traditional taxonomic approach is not sufficient to classify new species. In addition, the time and money needed for a morphological approach is much greater than that necessary for DNA barcoding. Only a small amount of tissue is required, even degraded (the whole organism is not necessary), in any develop stage (eggs, larvae, adults...). In this work a rapid random assessment sampling together with DNA barcoding tool have been used and emphasizes the value of DNA barcoding as a tool for biodiversity inventory (Ardura et al., 2011). This method is especially applicable to marine invertebrates, as there are many reference sequences available from many different geographical areas (Hebert et al., 2003; Bucklin et al., 2011; Geller et al., 2013). However, in some cases the database is 8
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not sufficient for assigning species as occurs with the three phenotypes from our sampling: they can be assigned at genus level (Table 2); in these cases a previous taxonomic work or sequencing of other genera is needed, but it is not the aim of this study, especially since they are not NIS and the genus is sufficient for biodiversity
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analysis. In this particular work, genetic barcoding helps to solve uncertainties, especially in ascidian species in which microscopic examination is usually necessary for species identification (Hirose & Hirose, 2013).
Both ecosystems sampled in this study have different natural and anthropogenic
characteristics. The Saint-Nazaire lagoon is less saline and has a lower temperature on
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average and it is involved in the Natura 2000 Network, however it presents serious eutrophication problems due to human activity (Souchu et al., 2010). On the other hand,
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Canet port sustains intense recreational boat traffic, without eutrophication problems but with urban pollution (Souchu et al., 2010). The diversity calculated based on different parameters (species richness, dominance and phylogenetic structure) is always greater in Canet port than in Saint-Nazaire lagoon (Table 3) and can be used as proxy biodiversity indicator (May, 1995; Bianchi & Morri, 2000; Katsanevakis et al. 2014). Three NIS species were found in Canet port while only one was found in Saint-Nazaire
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lagoon (Table 3). The presence of a greater number of NIS in Canet port is what would be expected of a busy harbor with significant maritime traffic (Hewitt & Martin, 1996; Firestone and Corbett, 2005; Leuven et al., 2009). On the other hand, the presence of Ficopomatus enigmaticus within the Saint-Nazaire Special Protection Area questions
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the efficacy of protected areas as a means of preventing the introduction of NIS. This introduced species was probably transported on the hulls of ships or on commercial mollusk shells (de Wit, 2011). It is most prominent and grows best in estuaries and
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lagoons with brackish waters and a high nutrient content (JNCC, 1997; Schwind & Iribarne, 2000; Bianchi & Morri, 2001; Paavola et al., 2005). The three NIS species present in Canet port, Crassostrea gigas, Aiptasia pulchella and Styela plicata, are native from the NW Pacific, probably originating from Japanese coasts. These species have already been described along French coasts (Gruet et al., 1976), and their introductions are most likely a result of the ‘controlled’ authorized introduction of C. gigas for aquaculture purposes from Japanese stocks (2001). The NIS identified in Saint-Nazaire Lagoon, Ficopomatus enigmaticus, is presumably native to Australia (McQuaid and Griffiths, 2014). It was first noticed in northern France in 1921 (de Wit, 9
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2011) and a previous study by Fornos et al. (1997) described its presence in the Western Mediterranean lagoons as well. It is probable that it was introduced either through transport on the hulls of ships or on commercial mollusk shells (de Wit, 2011). The reef-building polychaete F. enigmaticus has been shown to have major impacts on the
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systems it invades and is therefore considered an ecosystem engineer (Schwindt et al., 2001, 2004). Their impacts have been shown worldwide (i.e. Mar Chiquita coastal lagoon in Argentina (Schwindt et al., 2004) or in Zandvlei Estuary, Cape Town, South Africa (McQuaid and Griffiths, 2014)). This polychaete clearly impacts
native
biodiversity, its presence in the Saint-Nazaire lagoon represents to 20% of the total of
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biodiversity inventoried in this area (Table 3) this proportion being a signal of
significant biodiversity turnover (Katsanevakis et al., 2014), and it is higher than that
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found in Canet port (12.5%) corresponding to three different NIS.
Protected areas are a common tool for the conservation of biodiversity, vulnerable species, ecological processes and the protection of resources from overexploitation and from the destructive effects of human activities in both terrestrial and marine environments (Gaines et al., 2010; Pimm et al., 2001; Le Saout et al., 2013; Rife et al., 2013). The Natura 2000 network is currently the centerpiece of EU nature &
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biodiversity conservation policy, but as the data demonstrate that the mere establishment of Protected Areas does not guarantee their success. When protected areas are simply decreed but insufficient resources are available for effective design, management, or enforcement, these become only “paper parks” that do not effectively
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restrict exploitation or access (White & Courtney, 2004; Rife et al., 2013). For the Canet Saint-Nazaire Lagoon, this is the case, there is currently no management plan for the lagoon (see site code: FR9112025, Natura2000, 2016).
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The biggest concern is that estuaries have been urbanized worldwide as the popularity of coastal areas increases continuously (Chapman, 2006; Bockley, 2007) and their environmental protection is affected by economic interests. Natural ecosystems are being replaced by man-made ones with hard surfaces replacing soft sand or sediments (Bulleri and Airoldi, 2005; Chapman, 2006; Blockley, 2007; Glasby et al., 2007; Dafforn et al., 2009). These artificial structures are an ideal substrate for new coming species, normally dominated by NIS (Bulleri and Airoldi, 2005; Glasby et al., 2007). The newly arriving species, together with new artificial structures, are altering the habitat by increasing habitat complexity with an influence on the abundance, diversity 10
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and distribution of the organism chain (Coull and Wells, 1983) and native biodiversity is being threatened with habitat destruction and the introduction of NIS (Glasby et al., 2007). Therefore, habitat degradation is more of a promoter than a consequence of biological invasions (MacDougall and Turkington, 2005), and similar to terrestrial
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ecosystems, species richness of marine communities is negatively correlated with invasion success (Stachowicz et al., 1999). Thus, native biodiversity and habitat health
could be strong barriers for biological invasions. While tourism is a very important economic activity for many areas worldwide, some interventions can be carried out to
avoid the arrival and dispersion of new coming species, such as cleaning recreational
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vessels before transport between estuaries, avoiding the dumping of hard objects to limit the availability of suitable substrata for F. enigmaticus and other similar non-indigenous
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invertebrate species (McQuaid and Griffiths, 2004).
The methodology developed in this study is very valuable for a rapid assessment of invertebrate species biodiversity, because the individuals can be taken easily from the environment and it is a very fast, cheap and easy technique (without the need for very specialized human and technical resources) to inventory biodiversity and detect NIS. The problem is that sampling is still necessary and requires a large effort of human
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resources and many specialists systematically sampling all ecosystems, which is even greater in habitats with difficult physical access that have to be accessed from the sea and/or diving. In addition, some species cannot be detected because they are at a low density, in their first development stages or with higher mobility (not sessile species). In
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these cases, the use of environmental DNA, extracted directly from water samples, together with high-throughput sequencing (HTS) metabarcoding is capable of nontarget species detection and is highly sensitive (Ardura et al., 2016; Ardura et al.,
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2015b; Zaiko et al. 2015). However, the cost of the sequencing is significantly higher and substantial analytical effort (bioinformatics) is needed to ensure efficient exploration of the sequence data obtained from multi-species communities (Blanchet, 2012). For this approach taxonomic expertise is not required, eDNA-HTS techniques can supplement observational records and field surveys to obtain marine ecosystem samples and information, as they reveal the number of NIS in one specific area and their temporal and spatial distribution (Ardura et al., 2015; Zaiko et al., 2015). The best methodology should be assessed for each particular study, depending on economic and material resources and the data available and necessary in each case.
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Given the above, with one or another methodology, the early detection of NIS (target or not-target species), its knowledge about their distribution and abundance and the biodiversity inventory are essential to know the state and health of a particular environment. This knowledge is necessary for the management of NIS and it should go
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together with efforts and resources that nations have applied to reduce pollution and to restore wetlands and fishery stocks, as all of these elements are interconnected and they are promoters and consequences of the same problem: the anthropogenic impact on the
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environment.
Acknowledgments
Thank you to Martin Desmalades for helping with the sampling tasks and to Jeanine
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Almany for improvements on the manuscript. We are grateful to Thomas E. Holloway (Royal Holloway University of London, UK) for proofreading the manuscript. A.A. holds a regional postdoctoral Marie Curie grants COFUND-CLARIN. References
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Fig. 1. Map of the sampling points. Saint-Nazaire Lagoon and Canet port situated on the Mediterranean Sea, and the 5 points sampled of each location.
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Table 1. Sampling sites with GPS coordinates.
Latitude
Longitude
Saint-Nazaire-1
A1
42.67413° N
3.0134º E
Saint-Nazaire-2
A2
42.67641º N
3.00027º E
Saint-Nazaire-3
A3
42.67204º N
3.03250º E
Saint-nazaire-4
A4
42.65949º N
3.03439º E
Saint-Nazaire-5
A5
42.65261º N
Canet port-1
B1
42.70201º N
Canet port-3 Canet port-4
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B2
42.70433º N
3.03516º E
B3
42.70468º N
3.02776º E
B4
42.70414° N
3.03284° E
B5
42.70360° N
3.03845° E
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Canet port-5
3.03158º E
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Canet port-2
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Point
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Table 2. Species found from each sampling location, in percent (N= sample size). Nonindigenous species are marked in bold. GenBank Accesion Number (GB-AN).
Saint-Nazaire
Abra sp
Native
0.37
Aiptasia pulchella
NW Pacific
0.01
Anomia ephippium
Native
0.04
Balanus perforatus
Native
0.05
Carcinus aestuarii
Native
0.01
Cerastoderma glaucum
Native
Crassostrea gigas
NW Pacific
Ficopomatus enigmaticus
Australia
Gibbula adansonii
Native
Gibbula varia
Native
Hediste diversicolor
Native
Lumbrineris funchalensis
Native
Mimachlamys varia
16S GB-AN
-
KT988336
KT988316
-
KT988315
-
KT970486
-
-
KT988319
-
KT988318
-
-
KT988337
0.01
KT988320
-
0.05
KT988321
-
KT988322
-
0.01
-
KT988339
Native
0.01
KT988323
-
Native
0.05
-
KT988340
Native
0.29
KT988324
-
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Mytilus galloprovincialis
COI GB-AN
KT988317
0.01
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0.04
0.35
0.26
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Maja crispate
Canet port
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Status
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Species
Native
0.01
KT988325
-
Nassarius sp
Native
0.03
-
KT988341
Ocenebra erinaceus
Native
0.01
KT988327
-
Osilinus turbinatus
Native
0.04
KT988328
-
Ostrea edulis
Native
0.04
KT988326
-
Pachygrapsus marmoratus
Native
0.01
KT988331
-
Pagurus cuanensis
Native
0.01
KT988329
-
Patella caerulea
Native
0.18
KT988330
-
Patella rustica
Native
0.02
KT988332
-
Patella ulyssiponensis
Native
0.03
KT988333
-
Polynoe sp
Native
0.01
-
KT988342
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Nassarius incrassatus
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Sabella pavonina
Native
0.01
KT988334
-
Styela plicata
NW Pacific
0.04
KT988335
-
100
100
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N
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Table 3. Diversity indices values for invertebrate communities found in the two considered locations on the Mediterranean Sea. Number of species (NSp), species richness (SR), proportion of NIS (NIS), Simpson’s diversity index and phylogenetic diversity (PD).
SR (d)
NIS
Simpson
PD
Saint-Nazaire
5
0.869
0.200
0.680
416.7
Canet port
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4.994
0.125
0.882
1,433
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NSp
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