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Aquatic Invasions (2022) Volume 17, Issue 3: 308–3
52
Tsirintanis et al. (2022), Aquatic Invasions 17(3): 308–352, https://doi.org/10.3391/ai.2022.17.3.01 308
CORRECTED PROOF
Review
Bioinvasion impacts on biodiversity, ecosystem services, and human health
in the Mediterranean Sea
Konstantinos Tsirintanis1, Ernesto Azzurro2,3, Fabio Crocetta3, Margarita Dimiza4, Carlo Froglia2, Vasilis Gerovasileiou5,6,
Joachim Langeneck7, Giorgio Mancinelli8,9,10, Antonietta Rosso10,11, Nir Stern12, Maria Triantaphyllou4, Konstantinos
Tsiamis13, Xavier Turon14, Marc Verlaque15, Argyro Zenetos16 and Stelios Katsanevakis1,*
1Department of Marine Sciences, University of the Aegean, Lofos Panepistimiou, 81100 Mytilene, Greece; 2CNR-IRBIM, National Research Council.
Institute of Biological Resources and Marine Biotechnologies, Ancona, Italy; 3Department of Integrative Marine Ecology, Stazione Zoologica Anton
Dohrn, I-80121 Naples, Italy; 4Faculty of Geology & Geoenvironment, National and Kapodistrian University of Athens, Panepistimioupolis 15784,
Athens, Greece; 5Department of Environment, Faculty of Environment, Ionian University, 29100 Zakynthos, Greece; 6Hellenic Centre for Marine
Research (HCMR), Institute of Marine Biology, Biotechnology and Aquaculture (IMBBC), 71500 Heraklion, Greece; 7Department of Biology, University
of Pisa, 56126 Pisa, Italy; 8Department of Biological and Environmental Sciences and Technologies (DiSTeBA), University of Salento, SP Lecce-
Monteroni, 73100 Lecce, Italy; 9National Research Council – Institute of Marine Biological Resources and Biotechnologies (CNR-IRBIM), 71010
Lesina (FG), Italy; 10CoNISMa, Consorzio Nazionale Interuniversitario per le Scienze del Mare, 00196 Roma, Italy; 11Department of Biological,
Geological and Environmental Sciences, University of Catania, 95129 Catania, Italy; 12Israel Oceanographic and Limnological Research, National
Institute of Oceanography, Haifa 31080, Israel; 13Vroutou 12 Athens, 11141, Greece; 14Centre for Advanced Studies of Blanes (CEAB, CSIC), 17300
Blanes, Catalonia, Spain; 15Aix Marseille University and Université de Toulon, CNRS, IRD, Mediterranean Institute of Oceanography (MIO), UM 110,
Marseille, France; 16Institute of Marine Biological Resources & Inland Waters (IMBRIW), Hellenic Centre for Marine Research (HCMR), 16452,
Argyroupolis, Greece
*Corresponding author
E-mail: stelios@katsanevakis.com
Abstract
Biological invasions have become a defining feature of marine Mediterranean
ecosystems with significant impacts on biodiversity, ecosystem services, and human
health. We systematically reviewed the current knowledge on the impacts of marine
biological invasions in the Mediterranean Sea. We screened relevant literature and
applied a standardised framework that classifies mechanisms and magnitude of impacts
and type of evidence. Overall, 103 alien and cryptogenic species were analysed, 59 of
which were associated with both negative and positive impacts, 17 to only negative, and
13 to only positive; no impacts were found for 14 species. Evidence for most reported
impacts (52%) was of medium strength, but for 32% of impact reports evidence was
weak, based solely on expert judgement. Only 16% of the reported impacts were based
on experimental studies. Our assessment allowed us to create an inventory of 88 alien
and cryptogenic species from 16 different phyla with reported moderate to high
impacts. The ten worst invasive species in terms of reported negative impacts on
biodiversity include six algae, two fishes, and two molluscs, with the green alga Caulerpa
cylindracea ranking first. Negative impacts on biodiversity prevailed over positive ones.
Competition for resources, the creation of novel habitat through ecosystem engineering,
and predation were the primary reported mechanisms of negative effects. Most cases of
combined negative and positive impacts on biodiversity referred to community-level
modifications. Overall, more positive than negative impacts were reported on ecosystem
services, but this varied depending on the service. For human health, only negative
impacts were recorded. Substantial variation was found among Mediterranean ecoregions
in terms of mechanisms of impact and the taxonomic identity of impacting species.
There was no evidence that the magnitude of impact increases with residence time.
Holistic approaches and experimental research constitute the way forward to better
understanding and managing biological invasions.
Key words: alien species, cryptogenic, effects, experiments, expert judgement,
modelling, systematic review
Citation: Tsirintanis K, Azzurro E,
Crocetta
F, Dimiza M, Froglia C,
Gerovasileiou V, Langeneck J, Mancinelli G,
Rosso
A, Stern N, Triantaphyllou M,
Tsiamis K, Turon X, Verlaque M, Zenetos A,
Katsanevakis
S (2022) Bioinvasion
impacts on biodiversity, ecosystem
services, and human health in the
Mediterranean Sea
. Aquatic Invasions
17
(3): 308–352, https://doi.org/10.3391/ai.
2022.17.3.01
Received:
20 December 2021
Accepted:
19 May 2022
Published:
5 July 2022
Thematic editor:
Charles Martin
Copyright:
© Tsirintanis et al.
This is an open access article distributed under terms
of the Creative Commons Attribution License
(
Attribution 4.0 International - CC BY 4.0).
OPEN ACCESS.
Bioinvasions impacts in the Mediterranean Sea
Tsirintanis et al. (2022), Aquatic Invasions 17(3): 308–352, https://doi.org/10.3391/ai.2022.17.3.01 309
Introduction
The rate of human-mediated introduction of species to new areas is high
and accelerating all over the world (Seebens et al. 2017; Pyšek et al. 2020),
with heavy and rising associated economic costs (Diagne et al. 2021).
Modern biological invasions constitute a unique form of global environmental
change, profound enough to surpass natural drivers of selection and
dispersal, creating novel biological communities that would have never
emerged through natural processes (Ricciardi 2007). In the ecological
domain, the integration of novel species in the ecosystems can generate
fundamental alterations to the structure of native communities and
ecosystem functioning (e.g., Vitousek et al. 1996; Kideys 2002; Pranovi et
al. 2006; Galil 2007; Ehrenfeld 2010; Vilà et al. 2011; Katsanevakis et al.
2014a; Howard et al. 2019), often with detrimental losses of native
biodiversity (e.g., Clavero et al. 2009; Sala et al. 2011; Harper and Bunbury
2015; Bellard et al. 2016; Doherty et al. 2016; García-Gómez et al. 2020).
The provision of ecosystem services can also be strongly affected by the
impacts of biological invasions (Vilà et al. 2010; Katsanevakis et al. 2014a;
Vilà and Hulme 2017; Castro-Díez et al. 2019). Furthermore, invasive
species pose a threat to human health through a variety of mechanisms
such as pathogen transmission, intoxication, and envenomation (Schaffner
et al. 2013; Hulme 2014; Galil 2018; Peyton et al. 2019; Bédry et al. 2021).
Recent works highlighted that biological invasions can also have positive
outcomes (Katsanevakis et al. 2014a; Vimercati et al. 2020), for example by
creating novel ecosystems that may support native biodiversity and ecosystem
functioning (Hobbs et al. 2009; Schlaepfer et al. 2011; Evers et al. 2018),
providing ecosystem services (Katsanevakis et al. 2018; Apostolaki et al.
2019), and contributing to human well-being (Shackleton et al. 2019; Sfriso
et al. 2020). The positive impacts of alien species are dependent on societal
perception and different cultural values and motivations (Simberloff et al.
2013) and are commonly underestimated (Katsanevakis et al. 2014a).
Nevertheless, their study is receiving increasing attention (Vimercati et al.
2020) and, in some cases, alien species can even constitute targets for
protection and conservation (Schlaepfer et al. 2011; Mačić et al. 2018).
Impact assessment (including both negative and positive impacts of
biological invasions) is an important but challenging process aiming to
quantify the magnitude of changes and prioritise management and mitigation
of undesired effects (Parker et al. 1999; Kumschick et al. 2012; Ricciardi et
al. 2013; Simberloff et al. 2013; Ojaveer et al. 2015). The utilisation of
conceptual frameworks that classify in a standard way the effects of alien
species on ecosystems is essential in the process (Blackburn et al. 2011;
Bacher et al. 2018). Katsanevakis et al. (2014a) assessed the impacts of
marine alien species on biodiversity and ecosystem services utilising a
framework that emphasised the robustness of the reported evidence and
Bioinvasions impacts in the Mediterranean Sea
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Figure 1. Mediterranean Sea ecoregions (sensu Spalding et al. 2007). The present review covered
all seven Mediterranean ecoregions.
the ecological mechanisms of negative and positive impacts on both
biodiversity and ecosystem services. Blackburn et al. (2014) proposed a
classification framework of impact magnitude levels for a series of
ecological mechanisms through which alien species impact individuals,
populations, and communities. Difficulties that hinder impact assessments,
such as a wide range of taxonomic groups, various types of affected
habitats, estimation of impact magnitude, and spatial quantification of
impacts, can be overcome through the application of such schemes
(Kumschick et al. 2015; Katsanevakis et al. 2016; Nentwig et al. 2016;
Galanidi et al. 2018).
The Mediterranean Sea is a hotspot of biodiversity with a high level of
endemism (Coll et al. 2010). It is divided into seven marine ecoregions
(Figure 1), each with a distinct suite of oceanographic and topographic
features and relatively homogeneous species composition (Spalding et al.
2007). In addition, the Mediterranean is a hotspot of biological invasions,
particularly throughout its eastern part, having the highest number of
introduced species than any other sea region of the world (Costello et al.
2021). In the same easternmost sectors, under a rapid warming trend,
many thermally-sensitive native populations have collapsed in the past few
decades (Yeruham et al. 2015; Rilov 2016; Albano et al. 2021) and in some
cases alien species are the only ones that can sustain ecosystem functions.
Hence, a shift in conservation strategies from protecting native biodiversity
to protecting ecosystem functions (thus also including some alien species
as conservation targets) has been recently proposed (Katsanevakis et al.
2020a).
Approximately 1,000 species are estimated to have been introduced in
Mediterranean ecoregions (Zenetos et al. 2010, 2012), with more than half
Bioinvasions impacts in the Mediterranean Sea
Tsirintanis et al. (2022), Aquatic Invasions 17(3): 308–352, https://doi.org/10.3391/ai.2022.17.3.01 311
having established populations (Galil et al. 2016; Zenetos et al. 2017;
Zenetos and Galanidi 2020) and exhibiting an accelerating rate of
establishment success during recent years (Zenetos et al. 2022). Biological
invasions in the Mediterranean have caused large-scale biogeographic
modifications, range shifts of native species, population declines or even
local extinctions (Galil 2007; Edelist et al. 2013; Katsanevakis et al. 2014a).
Previous assessments of the spatial variation of impacts in the Mediterranean
Sea indicated strong spatial heterogeneity, not significantly correlated to
alien species richness (Katsanevakis et al. 2016). As the situation changes
fast, with many new alien species becoming invasive and expanding rapidly
in the last few years (e.g., Dimitriadis et al. 2020; Zenetos and Galanidi
2020), and as new knowledge accumulates, a regular reassessment of
impacts is important to adequately inform management measures.
With the purpose to update the current knowledge on invasive alien
marine species’ impacts on Mediterranean ecosystems, a new systematic
literature review was conducted building upon the work of Katsanevakis et al.
(2014a). To accomplish this, we compiled and assessed published available
information on alien and cryptogenic species impacts to: (i) identify
impactful species for biodiversity, ecosystem services, and human health,
updating previous knowledge, (ii) assess the magnitude and mechanisms of
their impacts, and their variability across Mediterranean Sea ecoregions,
and (iii) propose an inventory of the impactful marine alien species of the
Mediterranean Sea.
Materials and methods
List of assessed species
A list of targeted taxa for impact assessments was compiled, including
marine alien and cryptogenic species flagged as of “High Impact” in the
European Alien Species Information Network (EASIN; Katsanevakis et al.
2012) or as high-impact or invasive in recently published reviews (e.g.
Otero et al. 2013; Katsanevakis et al. 2014a; Karachle et al. 2017; Galanidi et
al. 2018; Zenetos et al. 2018; Peyton et al. 2019; Tsiamis et al. 2020). Based
on their expert knowledge, the authors added to the list further invasive
species not previously considered by EASIN or the aforementioned
reviews. Diatoms, dinoflagellates and other microalgae were not included
in the initial species list, based on Gómez (2008, 2019) who argued about
the difficulties and uncertainties in defining the native distribution of such
species. Tsiamis et al. (2021) have agreed that there are large gaps in
knowledge on unicellular plankton species and that more work is needed
on marine non-indigenous phytoplankton species in Europe before tagging
any phytoplankton species as alien. In total, 103 alien and cryptogenic
species belonging to 16 different taxonomic groups were assessed for their
impacts on biodiversity, ecosystem services, and human health in the
Mediterranean Sea.
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Literature review
A bibliographical review was performed for each targeted species, using the
Scholar Google search engine, as it includes grey literature (i.e. technical
reports, PhD and MSc theses, and conference proceedings) in addition to
peer-reviewed journal articles. Species previously assessed by Katsanevakis
et al. (2014a) were updated by searching for more recent publications
(between 2013 and March 2021), whereas the bibliographic search was
performed without chronological restrictions for those species not
previously assessed. For each species, eligible documents were those with
the following set of keywords (anywhere in the article): <species name>
AND “impact” AND “Mediterranean”. The search provided 62,043 articles
(sum of the articles found for each separate species search) but included
duplicates as some articles covered more than one target species. Potentially
eligible papers to be included in the review were initially screened based on
their titles and abstracts; for those that passed the initial screening, the full
text was reviewed. At a later stage, taxonomic experts further enriched the
list of selected studies with additional documents that had not been found
through the literature search. In total, 593 studies that included relevant
information on negative or positive impacts of alien and cryptogenic
species were retained for the analysis.
From each retained paper, the following information was extracted and
coded: DOI/link; short reference; year of publication; Mediterranean
ecoregion (according to Spalding et al. 2007; Figure 1); species name and
phylum following WoRMS (WoRMS Editorial Board 2021); and reported
negative or positive impact on biodiversity, ecosystem services, and human
health, type of evidence, mechanism of impact, magnitude of impact and
ecosystem engineer type, if relevant (the latter two only for impacts on
biodiversity). Only impacts referring to the Mediterranean Sea were
included; potential impacts, e.g. based on evidence in non-Mediterranean
regions, were not considered.
Framework for impact assessment
Type of evidence refers to the robustness of the methodological approach
that was applied to assess the impact. As such, the robustness of the reported
evidence was classified into six different types corresponding to three strength
of evidence categories (Katsanevakis et al. 2014a, 2016). The six evidence
types were: (i) manipulative experiments and (ii) natural experiments (high
strength of evidence), (iii) direct observations, (iv) modelling, (v) non–
experimental-based correlations (medium strength of evidence), and
(vi) expert judgement (low strength of evidence). Manipulative experiments
are field or laboratory experiments that include treatments/control and
random selection of experimental units. In natural experiments, the
experimental units (i.e. controls or impacted areas) are selected by nature
Bioinvasions impacts in the Mediterranean Sea
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Figure 2. Main mechanisms through which marine alien species impact biodiversity (sensu
Katsanevakis et al. 2014a). Red background refers to negative impacts and green background
refers to positive impacts.
(i.e. not randomly). Direct observations are direct raw measurements of
the impact about which there is no doubt (e.g. fouling in aquaculture or on
fishing gear, new commodities in fish markets). Modelling refers to
modelled consequences on biodiversity, ecosystem services and human
health derived from ecosystem models. Non-experimental-based correlations
refer to significant correlations between the targeted species and the
investigated impact, but not based on an experimental design for data
collection (e.g. there was a negative correlation between the biomass of an
alien predator and the biomass of its native prey). Expert judgement is a
qualitative or semi-quantitative assessment that relies on the empirical
knowledge of experts based on the species’ traits or the reported impact of
similar or the same species in a different geographical region.
Each reported impact was linked to the underlying mechanism.
Mechanisms of impact were classified according to Katsanevakis et al.
(2014a), with some additions (see Figure 2 for biodiversity; Figure 3 for
regulating and maintenance ecosystem services; Figure 4 for cultural,
provisioning ecosystem services, and human health). The impact magnitude
was classified following the five-level classification proposed by Blackburn
et al. (2014), i.e. minimal, minor, moderate, major, and massive (Table 1),
Bioinvasions impacts in the Mediterranean Sea
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Figure 3. Main mechanisms through which marine alien species impact regulating and
maintenance ecosystem services (sensu Liquete et al. 2013). Red background refers to negative
impacts while green background refers to positive impacts. The classification of impacts on
ecosystem services was based on Katsanevakis et al. (2014a).
considering also Volery et al. (2020). The selection of marine ecosystem
services classification was based on Liquete et al. (2013), who proposed a
classification of 14 marine ecosystem services grouped as “provisioning”,
“regulating and maintenance”, and “cultural” (Table 2). Ecosystem engineers
were categorised into four different types: structural, chemical, light engineers,
and bioturbators (Wallentinus and Nyberg 2007; Berke 2010; Katsanevakis
et al. 2014a; Figure 5).
Analyses
In addition to descriptive statistics, contingency table analyses were
conducted to quantify the degree of association between selected pairs of
variables. The hypothesis of independence between such pairs was tested
via chi-square tests. To avoid having cells with < 5 observations, certain
categories within each variable were grouped. Hierarchical cluster analysis,
using Euclidean distance and square-root transformation of the number of
impact records by each mechanism for each species, was performed to
determine the similarities of the analysed species in terms of the
mechanisms of their impacts for both negative and positive records.
Bioinvasions impacts in the Mediterranean Sea
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Figure 4. Main mechanisms through which marine alien species impact provisioning, cultural
ecosystem services (sensu Liquete et al. 2013) and human health. Red background refers to
negative impacts while green background refers to positive impacts. The classification of
impacts on ecosystem services was based on Katsanevakis et al. (2014a), with the exception of
categories marked with an asterisk (*) that are new.
To test the hypothesis that the magnitude of reported impacts on
biodiversity increases with the residence time, i.e. the number of years
passed since the first detection of the species in the Mediterranean Sea, we
followed two different approaches. In the first approach, a one-way
ANOVA was conducted with the maximum reported magnitude of impact
on biodiversity for each species as the categorical factor and the residence
time as the dependent variable. Levene’s test was applied to test homogeneity
of variances, and Tukey’s honestly significant difference (HSD) procedure
for multiple range tests. In the second approach, we created an impact
score in which for each species all reports of impacts were summed up in a
weighted sum with varying weights depending on the magnitude of impact:
massive 4; major 3; moderate 2; minor 1; minimal 0. The relationship
between the impact score and the residence time of each species was
investigated with a linear regression approach. All analyses were conducted
using Statgraphics Centurion XVI.
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Table 1. Impact magnitude classification levels with criteria description (adapted from Blackburn et al. 2014, considering also
Volery et al. 2020), referring to both negative and positive impacts.
Impact magnitude
Criteria description
Minimal
No effect on fitness of native species; negligible impact on native species due to competition, predation,
parasitism, toxicity, bio-
fouling, food provision or control of another invasive species; no effects on
ecosystem processes and ecosystem fun
ctioning; no chemical, physical or structural impact on the
ecosystem (not an ecosystem engineer).
Minor
Impact on individual of a native species due to competition, predation, parasitism, toxicity, bio-fouling, food
provision, or control of another inva
sive species without substantial population changes; minor impact on
ecosystem processes and ecosystem functioning with no related population changes; any environmental or
habitat alterations in chemical, physical or structural characteristics do not result in population changes.
Moderate
Native species population changes due to competition, predation, parasitism, toxicity, bio-fouling, food
provision or control of another invasive species but without changes in community composition; or impact
on
ecosystem processes and ecosystem functioning resulting in population declines/increases but no
substantial change in species composition; or ecological engineering, resulting in population
declines/increases but no substantial change in community composition.
Major
Changes in community composition through local or global extinction (negative) or re-establishment
(positive) of at least one native species, because of competition, predation, parasitism, toxicity, bio-fouling,
food provision or control of
another invasive species; impact on ecosystem processes and ecosystem
functioning resulting in species composition changes; or ecological engineering, resulting in change in
community composition. Major magnitude applies when under the hypothetical scenario of the alien species
extinction, the induced changes are considered as reversible within 10 years or within 3 generations of the
extinct/re-established native taxon/ -a, whichever is longer.
Massive
Same conditions as in “major” magnitude, but changes are considered as irreversible under the hypothetical
scenario of the alien species extinction, within 10 years or within 3 generations of the extinct/re-established
native taxon/-a.
Results
Impact matrix
The 593 studies on the 103 alien and cryptogenic species resulted in 1,343
records of negative and positive impacts on biodiversity, ecosystem
services, and human health (Table 3). A detailed description of impacts for
each species is given in the online Supplementary material Appendix 1.
Among the analysed taxa, Osteichthyes was the most numerous group
(24 species), followed by Mollusca (17 species), and Arthropoda
(Crustacea) (14 species), whereas 25 macroalgae were analysed, belonging
to Rhodophyta (12 species), Chlorophyta (8 species), and Ochrophyta
(5 species). Biodiversity was impacted by 70% of the studied species (72
taxa; Table 3), whereas the remaining ones (30%) showed no reported
impacts on biodiversity. Only 3% of the studied species had only positive
effects on biodiversity, whereas 27% had only negative ones. In most cases
(40%) the studied species impacted biodiversity both negatively and positively.
Among the 72 species impacting biodiversity, 38 represented ecosystem
engineers that caused alterations to the studied ecosystem as bioconstructors
(structural engineers), regulators of light penetration (light engineers),
modifiers of the chemical matrix of the environment (chemical engineers),
and bioturbators (Table 3). Most ecosystem engineers were macroalgae
(phyla Ochrophyta, Chlorophyta, and Rhodophyta) and Mollusca.
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Table 2. The applied marine ecosystem services classification scheme, with a description of each service (sensu Liquete et al. 2013).
Ecosystem service
Category
Description
Food provision
Provisioning
Provision of biomass from the marine environment for human
consumption. This includes all industrial, artisanal and
recreational fishing activities and aquaculture.
Water storage and provision
Provisioning
Provision of water for human consumption and other uses. In the
marine environment, these uses are mainly associated with coastal
lakes, deltaic aquifers, desalination plants, industrial cooling
processes, and coastal aquaculture in ponds and raceways.
Biotic materials and biofuels
Provisioning
Provision of biomass or biotic elements for non-food purposes,
including medicinal (e.g. drugs, cosmetics), ornamental (e.g.
corals, shells) and other commercial or industrial purposes, such
as fishmeal, algal or plant fertilisers, an
d biomass to produce
energy or biogas from decomposing material.
Water purification
Regulating and maintenance
Biochemical and physicochemical processes involved in the
removal of wastes and pollutants from the aquatic environment,
including treatment of
human waste, dilution, sedimentation,
trapping or sequestration (e.g. of pesticide residues or industrial
pollution); bioremediation; oxygenation of “dead zones”,
filtration and absorption; remineralisation; and decomposition.
Air quality regulation
Regulating and maintenance
Regulation of air pollutant concentrations in the lower
atmosphere.
Coastal protection
Regulating and maintenance
Natural protection of the coastal zone against inundation and
erosion from waves, storms or sea level rise by
biogenic and
geologic structures that disrupt water movement and thus stabilise
sediments or create protective buffer zones.
Climate regulation
Regulating and maintenance
The ocean acts as a sink for greenhouse and climate active gasses,
as inorganic
carbon is dissolved into the seawater and used by
marine organisms, a percentage of which is sequestered; perennial
large algae and higher plants can store carbon for longer periods.
Weather regulation
Regulating and maintenance
Influence on the local weather conditions, e.g. the influence of
coastal vegetation and wetlands on air moisture and, eventually,
on the saturation point and cloud formation.
Ocean nourishment
Regulating and maintenance
Natural cycling processes leading to the availability of nutrients in
seawater for the production of organic matter.
Lifecycle maintenance
Regulating and maintenance
The biological and physical support to facilitate the healthy and
diverse reproduction of species; this mainly refers to the
maintenance of key
habitats that act as nurseries, spawning areas
or migratory routes.
Biological regulation
Regulating and maintenance
Biological control of pests. The control of pathogens especially in
aquaculture installations, the role of cleaner fish in reefs,
biological control on the spread of vector borne human diseases,
and the control of invasive species.
Symbolic and aesthetic values
Cultural
This is about the exaltation of senses and emotions by seascapes,
habitats or species, and values put on coastal
natural and cultural
sites, and on the existence and beauty of charismatic habitats and
species such as corals or marine mammals.
Recreation and tourism
Cultural
Opportunities that the marine environment provides for relaxation
and entertainment,
including coastal activities such as bathing,
sunbathing, snorkelling, SCUBA diving, and offshore activities
such as sailing, recreational fishing, and whale watching.
Cognitive effects
Cultural
Inspiration for arts and applications (e.g. architectural designs
inspired by marine shells, medical applications replicating marine
organic compounds), material for research and education (e.g. as
test organisms for biological experiments), information and
awareness (e.g. respect for nature through the observati
on of
marine wild life).
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Figure 5. Ecosystem engineer types with impacts on biodiversity (both negative and positive).
Food provision (commercial and recreational fishing, aquaculture) was
the most impacted ecosystem service, with 33 species causing negative
impacts, 18 positive impacts, and 14 both (Table 3). Osteichthyes were the
taxonomic group including the highest number of species with an impact
on food provision (19 species). All impacts caused by macroalgae (in total
15 species) on this service were negative. The cultural ecosystem service of
“cognitive effects” had the second-highest number of impactful species,
with most of the studied taxa affecting this service positively, as they
constituted model species for biomonitoring or in vitro research studies.
Among regulating and maintenance services, “life-cycle maintenance” and
“climate regulation” were affected by the highest number of species.
Human health impacts were all negative, caused by 9 species.
Impacts on biodiversity
More negative than positive impacts on biodiversity were reported, with
468 records (78%) against 129 (22%) (Figure 6A). Caulerpa cylindracea was
the invasive species for which most negative impacts on biodiversity were
reported, with 53 cases. This chlorophyte of Australian origin thrives in a
variety of habitats, depths and environmental conditions, and negatively
affects species and communities in multiple ways (see Appendix 1). The
species is widespread in the Mediterranean and so are its effects on
biodiversity. Womersleyella setacea and Lophocladia lallemandii followed
in terms of negative impact records. These invasive rhodophytes form dense
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Table 3. Impact Matrix: Negative (red) and positive (green) impacts of marine alien and cryptogenic species on biodiversity and
ecosystem services of the Mediterranean. Colour indicates strength of evidence (vivid colours for high strength of evidence,
medium for medium strength, pale colours for low strength). For species without reported impacts or minimal impacts grey cells
are depicted for biodiversity and blank cells for ecosystem services and human health. Type of evidence is indicated for each
impact: Manipulative Experiments, E; Natural Experiments, N; Direct Observations of impact, O; Modelling, M; non-
experimental-based Correlations, C; Expert Judgement, J. (Cr): Cryptogenic species.
Impact on biodiversity
Impact on marine ecosystem services and human health
Provisioning Regulating and maintenance Cultural Health
Alien Spec ies
Minimal / No impact
Impact on species / communities
Ecosystem engineer
Food provision
Water Storage and provision
Biotic materials and biofuels
Water purification
Air quality regulation
Coastal protection
Climate regulation
Weather regulation
Ocean nourishment
Lifecycle maintenance
Biological regulation
Symbolic and aesthetic values
Recreation and tourism
Cognitive benefits
Human health
Cercozoa
Haplosporidium pinnae (Cr)
NOJ
Foraminifera
Amphistegina lobifera
CJ J O
J J
NJ
Ochrophy ta
Chrysonephos lewisii
O O
O O
Rugulopteryx okamurae
NOCJ NO N O
O
NO
O O O
Sargassum muticum
J OCJ O O J J J
J
J J
O O J
EJ
Stypopodium schimperi
OJ OJ O
J
O OJ EJ
Undaria pinnatifida
J O O O
J
J
O EJ
Chlorop hyta
Caulerpa cyl indracea
ENOMCJ NOCJ ENOJ
J
J
EJ
J ENOM J
EJ
J J J
EOCJ
Caulerpa tax ifolia
ENOJ ENOJ NO OJ
J
J EJ
NJ
J J J
J
Caulerpa taxifolia var. distichophyl la
EJ N N O
J
Cladophora patentiramea
O O
Codium arabicum
J
O O
Codium fragile subsp. fragile
OJ O OJ EO O
J
J
J
J
O O J
C
Codium parvu lum
J
O
O O
Halimeda incr assata
N N N
Rhodophy ta
Acrothamni on preissii
ENOCJ OJ O O
J
J J J
Agarophyton vermiculophyllum
CJ NOJ NO O
J
J
O O O J
CJ
Antithamni on nipponicum
J
O
Asparagopsis armata
EJ O OCJ OC O
J
J J J
EJ
Asparagopsis taxiformis
EOCJ N
J J
EJ
Bonnemaisonia hamifera
Galaxaura ru gosa
J
O O
Ganonema farinosum (Cr)
Grateloupia turuturu
J OJ O O
Lophocladia lallemandii
ENOCJ NOCJ NO J
J
J J
J J
OJ J J
Polysiphonia morrowii
Womersleyella setacea
ENOCJ NOC OC OJ
J J
J J
J ENOJ
J J J
Tracheophyta
Halophila stipulacea
C O O OC J
J
J NJ
J J
C
Porifera
Paraleucilla magna
EJ
O
Cnidaria
Macrorhync hia philippina
OJ
Oculina patagonica (Cr)
OCJ O O J
J
J
J J J
E
Rhopilema nomadica
OJ E O O
O OCJ CJ OJ
Ctenophora
Mnemiopsis leidyi
ENC J
O
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Table 3. (continued).
Bryozoa
Amathia ver ticillata (Cr)
E O O O
Tricellaria inopinata
J
OJ
Mollusca
Anadara kagoshimensis
J OJ OJ J
J
J J
Anadara transversa
ECJ OJ OJ EOJ
J
J J
Arcuatula senhousia
EOCJ EOCJ EOCJ
OJ
J
O J
EJ
Brachidontes pharaonis
ECJ ECJ
NOCJ OCJ
O
J
J
J O
Bursatell a leachii (Cr)
O
E
Chama pacific a
J OJ OJ
O J
J
J J
CJ
Conomurex p ersicus
J C O
J
EC
Crepidula fornicata
Dendostrea cf. folium
Fulvia fragilis
J
J
N
Magallana/Crassostrea species
EOJ OJ OCJ O
EJ
ECJ
J J
J J J
J J J J J
Mya arenaria
Petricolaria pholadiformis
Pinctada radiata
O O O O O
J
J
J J
EC
Rapana venosa
O
Ruditapes philippinarum
CJ ENOCJ EOCJ
O
MJ
M MJ
J
ENJ
Spondylus spinosus
CJ OJ OJ O
O J
J
J J
NC
Annelida: Polychaeta
Ficopomatus enigmaticus
O OJ
EN
Hydroides e legans
EOJ EO EO O
O
Arthropoda: Crustacea
Acartia (Acanthacartia) tonsa (Cr)
C
N
Callinectes sapidus
EOCJ
OJ O
CJ
Dyspanope us sayi
E
Erugosquilla massavensis
CJ
OJ O
E
Matuta victor
Metapenaeus monoceros
MCJ M
CJ O
EC
Metapenaeus stebbingi
O
O
E
Paracerceis sculpta
J
Penaeus azte cus
O
Penaeus pulchricaudatus
OMJ M
J O
EC
Penaeus semisulcatus
OM M
O
NCJ
Percnon gibbe si (Cr)
O
Portunus segnis
OJ
OC O
EC
Rhithropanopeus harrisii
Echinodermata
Diadema setosum
J
Synaptula reciprocans
Chordata: Ascidi acea
Botrylloides violaceus
Botrylloides diegensis
Ciona robusta
O
EC
Clavelina obl onga
O
Didemnum vexillum
O
O
Herdmania m omus
CJ
Microcosm us squamiger
OC C
O
Polyandroc arpa zorritens is
E
E
Styela plicata
E
OJ
E
E
ENC
Chordata: Osteichthyes
Alepes djedaba
O
Apogonichthyoides pharaonis
Atherinomorus forskalii
OJ
O
Decapterus russelli
O
Dussumieria elopsoides
O
Etrumeus golanii
O
Fistularia c ommersonii
NOM M
MJ OM
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Table 3. (continued).
Herklotsic hthys punctatus
O
Lagocephalus sceleratus
OMJ
OMJ
OJ
Nemipterus randalli
CJ
J O
NC
Parexocoetus mento
Parupeneus forsskali
C
C O
Pempheris rhomboidea
OMJ J J
Plotosus lineatus
CJ
O
OJ
Pterois miles
O
J
OJ
Sargocentron rubrum
OC J J O
Saurida lesse psianus
MJ
J O
N
Scomberom orus commers on
O
Siganus luridus
ENMCJ
MJ J J J O
J J
J
EJ
J J J OJ
Siganus rivulatus
ENMCJ
MJ J J J O
J J
J
EJ
J J J J
Sphyraena c hrysotaenia
C
O
Torquigener flavimaculos us
J
Upeneus moluccensis
NC
C O
Upeneus pori
NC
C O
Figure 6. A) Number of impact records that resulted from the review analysis to impact
biodiversity. Number of impact records by mechanism through which alien species impacted
biodiversity B) negatively or/ and C) positively.
algal turfs that overgrow other species and alter ecologically important
habitats, with negative implications in benthic communities. Although
widespread in the Mediterranean, their effects were mainly reported from
the western Mediterranean ecoregion. Apart from macroalgae, the mollusc
Brachidontes pharaonis and the fishes Siganus luridus and Siganus rivulatus
were the species with most negative impact records among the studied
taxa. The Erythrean B. pharaonis is an old invader of the Mediterranean,
with a wide range of distribution, but it did not exhibit an alarming
invasive behaviour until the late 1990s. Subsequently, its populations
exploded in many eastern Mediterranean localities, becoming dominant in
shallow rocky reef platforms with extremely high densities, outcompeting
native species and shaping community composition (see Appendix 1). The
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Figure 7. Mosaic chart of mechanisms of negative impacts on biodiversity for the studied
different phyla.
Lessepsian invaders Siganus spp. dominate fish communities in many parts
of the eastern Mediterranean and have caused detrimental losses on
Mediterranean macroalgal communities through their herbivory (see
Appendix 1 for more information).
Caulerpa cylindracea was also the species most reported for its positive
impacts, with 12 records, followed by B. pharaonis and the serpulid
Ficopomatus enigmaticus. Although the dense populations of C. cylindracea
constitute a nuisance for the majority of native biota, it can also facilitate
several other species, from seagrasses to invertebrates. Brachidontes pharaonis
has acquired a functional role in Mediterranean ecosystems and contributes
significantly as a structural ecosystem engineer, as a filter-feeder that
reduces turbidity and increases light penetration, and as a food source for
native biota. Ficopomatus enigmaticus is a reef-building structural engineer
that provides novel habitat for many benthic taxa.
Negative impacts were caused by a variety of mechanisms (Figure 6B, C;
Figure 7), significantly differing in frequency by taxonomic group (chi-square
test; keeping the three most frequent mechanisms and grouping all others;
keeping the six most frequent taxonomic groups and grouping all others;
p < 0.001). Competition for resources was the most reported mechanism of
negative impacts with almost 250 records, followed by the creation of novel
habitat, and predation (Figure 6B). This was the case for most studied
phyla except Chordata (Osteichthyes), Annelida (Polychaeta), and Cercozoa
(Figure 7). Osteichthyes caused negative impacts mostly by preying on
native biota. Hydroides elegans and F. enigmaticus, the two studied alien
polychaetes, mainly affected other species through the creation of novel
Bioinvasions impacts in the Mediterranean Sea
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Figure 8. Geographical variation of the A) negative and B) positive mechanisms of impact on
biodiversity by Mediterranean ecoregion and of analyzed alien phyla with C) negative and D)
positive impacts on biodiversity per Mediterranean ecoregion. The size of the pie charts reflects
the number of reported cases (the inset indicates the minimum and maximum sizes).
habitat. This was also an important mechanism of negative impacts for
macroalgae such as L. lallemandii, C. cylindracea, Caulerpa taxifolia and
molluscs such as B. pharaonis, and Spondylus spinosus. Haplosporidium
pinnae, the sole representative of Cercozoa, has been associated with the
disease that has caused the extinction of the endemic bivalve Pinna nobilis
from almost the entire Mediterranean Sea, except for a small number of
refugia. The creation of novel habitat was the most common mechanism of
positive impacts on biodiversity accounting for 65% of the reported
positive cases (Figure 6C). Species such as F. enigmaticus, S. spinosus, the
bivalves of the genus Magallana/Crassostrea (see the notes in the Appendix 1
for the taxonomic status of the genus), and the rhodophyte Asparagopsis
armata create novel habitats that support a variety of species.
Analysis of impacts by ecoregion revealed that interspecific competition
is the most important mechanism generating negative impacts on biodiversity
in all Mediterranean ecoregions (Figure 8A); nevertheless, the frequency of
the various mechanisms significantly differed by ecoregion (chi-square
test; keeping the three most frequent mechanisms—competition, creation
of novel habitat, predation—and grouping all others; p < 0.001). The creation
of novel habitat was the second most reported negative impact mechanism
in western and central Mediterranean ecoregions. On the contrary, predation
was the second most impactful mechanism in the eastern Mediterranean
ecoregions (Figure 8A). The frequency of the mechanisms of reported
positive impacts also differed by ecoregion (chi-square test; keeping the
two most frequent mechanisms and grouping all others; excluding ecoregions
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with < 5 records; p = 0.008). Structural ecosystem engineers that create
novel habitats were the main contributors to positive impacts in all
Mediterranean ecoregions, except for the Levantine Sea where positive
impacts of food provision slightly exceeded the creation of novel habitat
(Figure 8B). Food provision to native biodiversity was also important in
the western Mediterranean and the Ionian Sea (Figure 8B).
There was significant variation in the taxonomic composition of the
species that caused negative impacts on biodiversity among ecoregions
(chi-square test; keeping the five most reported phyla and grouping all
others; p < 0.001) (Figure 8C). In the western Mediterranean, the most
reported negative impacts were caused by invasive Rhodophyta and
Chlorophyta, whereas in the Levantine and the Aegean Sea by Osteichthyes,
and in the Adriatic Sea by Mollusca (Figure 8C). Caulerpa cylindracea, L.
lallemandii, W. setacea, C. taxifolia, and Acrothamnion preissii were among
the chlorophytes and rhodophytes with negative effects in the western
Mediterranean whereas, in the Levantine Sea, B. pharaonis, S. luridus and
S. rivulatus were the species with the most recorded impacts. Siganus
luridus was also the species that accounted for most of the recorded
negative impacts in the Aegean Sea. Negative impacts caused by Crustacea
had an important share of total impacts in the Tunisian plateau/Gulf of
Sidra, the Levantine, and the Ionian Sea, but they had little relevance in the
western Mediterranean and the Alboran Sea (Figure 8C). The Alboran Sea
has been plagued by the negative impacts of the brown alga Rugulopteryx
okamurae, a recent invader that has caused detrimental effects on local
ecosystems (see Appendix 1 for more information).
There was also significant variation in the taxonomic composition of the
species that caused positive impacts on biodiversity among ecoregions
(chi-square test; keeping only the three ecoregions with the highest
number of records and the four mostly reported phyla, grouping all others;
p = 0.01). Macroalgae constituted the bulk of positive impact records for
the western Mediterranean, the Alboran Sea, and the Tunisian plateau/Gulf
of Sidra and had an important share in total reported positive impacts
from the Adriatic and the Ionian Sea (Figure 8D). Mollusca ranked first in
terms of reported positive impacts in the Levantine and the Adriatic Sea
and had an important share in the western Mediterranean and the Aegean
Sea. As creators of novel habitats, Polychaeta contributed to positive
impacts in the Aegean, Ionian, Adriatic Sea, and western Mediterranean
(Figure 8D). Furthermore, some Osteichthyes were reported to have positive
impacts on biodiversity, because of their role in the novel trophic webs, but
only in the Levantine Sea.
“Major” was the most reported impact magnitude category for both
negative and positive impact records (Figure 9). Nevertheless, the frequency
of impact magnitudes substantially differed between negative and positive
impacts, with “major” impacts having a significantly higher frequency in
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Figure 9. Proportion of impact magnitude categories of negative (left column) and positive
(right column) impacts on biodiversity.
positive than negative impacts (chi-square test; excluding minimal and
massive magnitude categories due to their low frequencies; p = 0.011).
Overall, 61% of negative impacts and 74% of positive impacts were
classified as major; no positive impacts of massive magnitude were reported.
The frequency of the categories of magnitude of impact significantly
differed among taxonomic groups (chi-square test; disregarding minimal
and massive magnitude categories due to their low frequencies, combining
phyla with a small number of records together; p < 0.001). Major negative
impacts were the primary magnitude category recorded for Rhodophyta,
Chlorophyta, Ochrophyta, Mollusca, Polychaeta, Cnidaria, Foraminifera,
Ascidiacea, and Bryozoa (Figure 10A). Massive negative impacts were only
recorded for Osteichthyes (Siganus spp.), Cercozoa (H. pinnae), and
Ochrophyta (R. okamurae) (Figure 10A). Moderate and minor impacts
were the most reported for Crustacea. Although the number of records of
positive impacts was lower than that of negative impacts, they followed a
similar pattern regarding impact magnitude (Figure 10B), with significant
differences among taxonomic groups (chi-square test; combining phyla
with a small number of records together; p = 0.011).
There was no evidence that the magnitude of impact increases with
residence time. The average residence time of introduced species did not
differ by the maximum reported magnitude of negative (ANOVA; p = 0.947)
or positive (ANOVA; p = 0.300) impacts. The slope of the linear regression
between the residence time of introduced species and the negative impact
score did not differ significantly from zero (p = 0.985), and the intercept
(i.e. the average score ± S.E.) was 11.3 ± 3.3 (Figure 11). The slope remained
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Figure 10 . Impact magnitude of A) negative and B) positive impacts on biodiversity per
phylum of alien species analysed..
Figure 11. Regressions of negative (left) or positive (right) impact scores in relation to residence time (years since first detection)
for each of the assessed species.
statistically not different from zero even when three influential points
(with leverage values greater than five times that of an average data point)
were removed (p = 0.529). Similarly, the slope of the linear regression
between the residence time of introduced species and the positive impact
score did not significantly differ from zero (p = 0.215), and the intercept
(i.e. the average score ± S.E.) was 2.2 ± 1.0 (Figure 11). The slope remained
statistically not different from zero even when three influential points
(with leverage values greater than five times that of an average data point)
were removed (p = 0.057).
Bioinvasions impacts in the Mediterranean Sea
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Figure 12 . Overview of negative and positive impacts on ecosystem services and human health.
Impacts on ecosystem services
Overall, 685 negative and positive impact records were found to affect
provisioning, regulating/maintenance, and cultural ecosystem services,
with 312 negative and 373 positive records in total, whereas for human
health 61 negative and no positive impacts were recorded. Most impact
records on provisioning services were recorded for food provision, with
116 negative and 188 positive records (Figure 12). Negative impacts prevailed
over positive for regulating/maintenance services. Still, for several
regulating and maintenance services (climate regulation, water purification,
coastal protection, and biological regulation) more positive than negative
impacts were reported (Figure 12). Among regulating/maintenance services,
lifecycle maintenance was the one with the highest number of recorded
negative impacts. No impact records were found for weather regulation.
For the cultural services “aesthetic values” and “recreation and tourism”,
70 negative and only 6 positive impact records were reported. On the other
hand, for “cognitive effects” records, positive impacts prevailed by 114
records against only 12 cases of negative impacts.
The frequency of the mechanisms of reported negative impacts on
provisioning ecosystem services significantly differed by ecoregion (chi-
square test; keeping the four most frequent mechanisms and excluding the
Alboran Sea; p < 0.001). Fouling shellfish, gear, and equipment of aquaculture
facilities was the primary mechanism of negative impacts on provisioning
services in the Adriatic Sea, the Ionian Sea, and the western Mediterranean
(Figure 13A). Entanglement in nets was the most important mechanism in
the Levantine and the Alboran Sea, with a considerable contribution to
negative impacts on provisioning services throughout the Mediterranean
Bioinvasions impacts in the Mediterranean Sea
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Figure 13 . Geographical variation of the A) negative and B) positive mechanisms of impact on
provisioning ecosystem services by Mediterranean ecoregion and of analyzed alien species
phyla with C) negative and D) positive impacts on provisioning services per Mediterranean
ecoregion. The size of the pie charts reflects the number of reported cases (the inset indicates
the minimum and maximum sizes).
(Figure 13A). Predation was also a primary mechanism of negative
impacts, especially in the eastern Mediterranean. From the analysis of
positive impacts by ecoregion on provisioning services (Figure 13B), it
became obvious that the Levantine Sea was the main beneficiary, due to
new commodities for local fisheries. Most positive impacts were recorded
in this ecoregion and significantly declined westwards (Figure 13B).
Among all mechanisms positively impacting provisioning services across
Mediterranean ecoregions, “new commodities” were by far the most
important.
The taxonomic synthesis of the species that negatively impacted
provisioning services differed significantly among ecoregions (chi-square
test; keeping the four most frequent phyla and merging all the others;
excluding the Alboran Sea; p < 0.001) (Figure 13C). Ascidians were the
major group of negative impacts in the western Mediterranean, with an
important contribution also in the Adriatic Sea (Figure 13C). This is due to
species such as the alien ascidians Styela plicata and Clavelina oblonga that
commonly foul farmed shellfish and aquaculture equipment. Molluscs
(Pinctada radiata), bryozoans (Tricellaria inopinata), and sponges
(Paraleucilla magna) also foul aquaculture facilities, introducing negative
effects on production. Cnidaria and Osteichthyes were the main groups
that impacted provisioning services in the Levantine Sea (Figure 13C). The
invasive jellyfish Rhopilema nomadica formed massive swarms that became
a nuisance for the local fisheries due to entanglement in nets and for water
Bioinvasions impacts in the Mediterranean Sea
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supply for local power plants and desalination facilities due to blocking of
intake pipes. The fish Lagocephalus sceleratus was also a major pest for
fisheries in the Levantine and the Aegean Sea due to its predation on the
catch and the damage caused to fishing gear. Alien crustacean species were
also reported to negatively impact the provision of food across multiple
ecoregions (Figure 13C). The alien crabs Callinectes sapidus and Portunus
segnis interfered with local fisheries in multiple Mediterranean ecoregions
by preying on the catches and by getting entangled in nets. Alien shrimps
such as Metapenaeus monoceros and Penaeus pulchricaudatus have been
associated with population declines of the commercially important native
shrimp Penaeus kerathurus (see Appendix 1).
The main contributors of positive impacts on food provision were alien
Osteichthyes and Crustacea, with positive impact records substantially
increasing eastwards (Figure 13D). Several alien fish such as Nemipterus
randalli, Scomberomorus commerson, Sphyraena chrysotaenia, Saurida
lessepsianus, Upeneus spp., Siganus spp. and crustaceans such as
P. pulchricaudatus, and Metapenaeus spp. are marketed in many
Mediterranean countries. Callinectes sapidus is collected and sold in the
northern Adriatic and Ebro Delta (Spain) where it is an appreciated shellfish.
Mollusc species, such as those of the genus Magallana/Crassostrea, Ruditapes
philippinarum, and Conomurex persicus have historically contributed to
food provision as new commodities in various Mediterranean ecoregions.
Degradation of habitats was the primary mechanism of negative impacts
on regulating services, mainly recorded in the western Mediterranean and
affecting life-cycle maintenance (Figure 14A). Posidonia oceanica meadows
and coralligenous communities are marine habitats with a fundamental
role in the functioning of Mediterranean ecosystems and the conservation
of marine life. Invasive Rhodophyta such as W. setacea, A. preissii, and
L. lallemandii have the potential to degrade such ecosystems (Figure 14C)
through the formation of dense algal turfs that epiphytise Posidonia
meadows and homogenise the diversity and complexity of coralligenous
communities, degrading their state and negatively affecting the life they
support (see Appendix 1). Such habitats are also vulnerable to the invasion
by C. cylindracea that negatively affects many keystone species. Furthermore,
the recent Ochrophyta invader R. okamurae can displace native species,
monopolise shallow communities, and even cause community shifts in
coralligenous communities.
Positive mechanisms of impact on “regulating and maintenance” ecosystem
services were much less reported (Figure 14B). Carbon sequestration
overall accounted for most positive impacts either through the formation
of shells or via primary production. Such positive contributions to climate
change were attributed to Mollusca, Foraminifera, and Tracheophyta
(Figure 14D). Shell-forming organisms such as the foraminiferan Amphistegina
Bioinvasions impacts in the Mediterranean Sea
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Figure 14 . Geographical variation of the A) negative and B) positive mechanisms of impact on
regulating and maintenance ecosystem services by Mediterranean ecoregion and of analyzed
alien species phyla with C) negative and D) positive impacts on regulating and maintenance
services per Mediterranean ecoregion. The size of the pie charts reflects the number of reported
cases (the inset indicates the minimum and maximum sizes).
lobifera or the bivalves R. philippinarum and Arcuatula senhousia form
dense aggregations that constitute banks of carbon storage in the marine
environment, since their shells contain carbonate. Meadows of the alien
seagrass Halophila stipulacea are also considered carbon sinks (see
Appendix 1).
The Levantine Sea was the Mediterranean ecoregion with most studies
of impacts on cultural ecosystem services (Figure 15A, B). The negative
effects in the Levantine were primarily caused by Chlorophyta, Ochrophyta,
and Rhodophyta (Figure 15C) through decomposing macroalgal material
that washed ashore. Seaweed drifts of Codium parvulum, Codium arabicum,
Galaxaura rugosa, and Stypopodium schimperi have regularly occurred in
the eastern Mediterranean during recent decades, disturbing local activities
and tourism and degrading the aesthetic value of the coast. Large and
dense macroalgal thalli of Sargassum muticum and Undaria pinnatifida
have sometimes become a nuisance in Adriatic lagoons, hindering navigation.
In the Levantine Sea, R. nomadica massive swarms have also hindered
tourism profits as they pose a threat to swimmers and deter tourists from
visiting the coast.
On the other hand, alien species have offered materials for research on
potential pharmaceutical products and biomonitoring services, which are
the main positive impacts reported on cultural services (Figure 15B), with
frequencies that were not found to differ significantly among ecoregions
(chi-square test; excluding the Alboran Sea; p = 0.136). Many molluscs and
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Figure 15 . Geographical variation of the A) negative and B) positive mechanisms of impact on
cultural ecosystem services by Mediterranean ecoregion and of analyzed alien species phyla
with C) negative and D) positive impacts on cultural services per Mediterranean ecoregion. The
size of the pie charts reflects the number of reported cases (the inset indicates the minimum and
maximum sizes).
crustaceans such as S. spinosus, P. radiata, and Penaeus semisulcatus, and
the foraminiferan A. lobifera, have been used effectively as biomonitoring
species for marine pollution (Figure 15D; see Appendix 1).
Impacts on human health
Reported impacts on human health were only negative. The great majority
of this evidence came from the Levantine Sea (Figure 16A) and was primarily
related to stinging and poisoning/intoxication caused by only 9 species
belonging to Osteichthyes, Cnidaria, and Echinodermata (Figure 16B),
namely: L. sceleratus, Torquigener flavimaculosus, Plotosus lineatus, P. miles,
S. luridus, S. rivulatus, Macrorhynchia philippina, R. nomadica, and Diadema
setosum. Such unwanted accidents are sometimes described in the scientific
literature. For example, in the eastern Mediterranean, severe intoxications
and cases of death were attributed to the consumption of L. sceleratus,
which contains in its tissues a paralytic neurotoxin called tetrodotoxin that can
be lethal for humans. Rhopilema nomadica has caused a vast number of
hospitalizations of swimmers and fishers in the Levantine Sea due to its painful
stings. The long-spined sea urchin D. setosum poses a threat for swimmers
in the eastern Mediterranean where it can be found at high densities in
shallow waters, as its spines contain venom and may inflict painful stings.
Grouping of species according to the mechanisms of their impacts
Species with impacts on biodiversity, ecosystem services, and human
health were grouped by hierarchical cluster analysis according to their
Bioinvasions impacts in the Mediterranean Sea
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Figure 16 . Geographical variation of A) mechanisms of negative impact and the responsible B)
phyla causing the impacts on human health by Mediterranean ecoregion.
Figure 17. Hierarchical cluster dendrogram of alien and cryptogenic species (Cr: cryptogenic) based on the mechanisms of
reported impacts on biodiversity, ecosystem services and human health. Coloured frames (red, green, blue, light blue) indicate the
four major groups species of the dendrogram. Included species are numbered as follows: 1) Acartia (Acanthacartia) tonsa (Cr), 2)
Acrothamnion preissii, 3) Agarophyton vermiculophyllum, 4) Alepes djedaba, 5) Amathia verticillata (Cr), 6) Amphistegina
lobifera, 7) Anadara kagoshimensis, 8) Anadara transversa, 9) Antithamnion nipponicum, 10) Arcuatula senhousia, 11) Asparagopsis
armata , 12) Asparagopsis taxiformis, 13) Atherinomorus forskalii, 14) Brachidontes pharaonis, 15) Bursatella leachii (Cr),
16) Callinectes sapidus, 17) Caulerpa cylindracea, 18) Caulerpa taxifolia, 19) Caulerpa taxifolia var. distichophylla, 20) Chama
pacifica, 21) Chrysonephos lewisii, 22) Ciona robusta, 23) Cladophora patentiramea, 24) Clavelina oblonga, 25) Codium
arabicum, 26) Codium fragile subsp. fragile, 27) Codium parvulum, 28) Conomurex persicus, 29) Decapterus russelli,
30) Diadema setosum, 31) Didemnum vexillum, 32) Dussumieria elopsoides, 33) Dyspanopeus sayi, 34) Erugosquilla massavensis,
35) Etrumeus golanii, 36) Ficopomatus enigmaticus, 37) Fistularia commersonii, 38) Fulvia fragilis, 39) Galaxaura rugosa,
40) Grateloupia turuturu, 41) Halimeda incrassata, 42) Halophila stipulacea, 43) Haplosporidium pinnae (Cr), 44) Herdmania
momus, 45) Herklotsichthys punctatus, 46) Hydroides elegans, 47) Lagocephalus sceleratus, 48) Lophocladia lallemandii,
49) Macrorhynchia philippina, 50) Magallana/Crassostrea species, 51) Metapenaeus monoceros, 52) Metapenaeus stebbingi,
53) Microcosmus squamiger, 54) Mnemiopsis leidyi, 55) Nemipterus randalli, 56) Oculina patagonica (Cr), 57) Paracerceis
sculpta, 58) Paraleucilla magna, 59) Parupeneus forsskali, 60) Pempheris rhomboidea, 61) Penaeus aztecus, 62) Penaeus
pulchricaudatus, 63) Penaeus semisulcatus, 64) Percnon gibbesi (Cr), 65) Pinctada radiata, 66) Plotosus lineatus,
67) Polyandrocarpa zorritensis, 68) Portunus segnis, 69) Pterois miles, 70) Rapana venosa, 71) Rhopilema nomadica, 72) Ruditapes
philippinarum, 73) Rugulopteryx okamurae, 74) Sargassum muticum, 75) Sargocentron rubrum, 76) Saurida lessepsianus,
77) Scomberomorus commerson, 78) Siganus luridus, 79) Siganus rivulatus, 80) Sphyraena chrysotaenia, 81) Spondylus spinosus,
82) Styela plicata, 83) Stypopodium schimperi, 84) Torquigener flavimaculosus, 85) Tricellaria inopinata, 86) Undaria
pinnatifida, 87) Upeneus moluccensis, 88) Upeneus pori, 89) Womersleyella setacea.
similarity of mechanisms of reported impacts, resulting in 4 major groups
(Figure 17).
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The first cluster (red colour in Figure 17) included six alien macroalgae,
namely C. cylindracea, W. setacea, L. lallemandii, R. okamurae, A. preissii,
and C. taxifolia. A high number of negative impact records was reported
for each of these species, with competition for resources being the most
common mechanism of impact. A high number of impact records was also
reported on ecosystem services through the mechanism of degradation of
important habitats, with 84 negative cases on 8 different services. These
species also constitute ecosystem engineers that mainly impacted other
species negatively (34 records) rather than positively (21 records) through
the creation of novel habitat.
The second cluster (green colour in Figure 17) was formed by 26 species
that were grouped together primarily due to their positive impact on food
provision, being new commodities. All these species, mainly crustaceans,
fishes, and two molluscs, contributed to 91% of the overall positive impacts
on food provision by new commodities. Among them, the alien shrimps
M. monoceros, P. pulchricaudatus, and P. semisulcatus were the species
with the highest number of positive reports as new commodities, but also
some competitive negative impacts on native shrimp species and thus
exhibited a high similarity. The alien fishes N. randalli, Saurida lessepsianus,
Upeneus spp., the alien molluscs C. persicus and R. philippinarum, and the
alien crab P. segnis formed a subgroup within this cluster following the
same pattern of reported mechanisms of impact but with a lower
frequency. Predation (including grazing) was also a mechanism of negative
impact characteristic for some species of this cluster, such as the
herbivorous fish Siganus spp., the omnivorous crab C. sapidus, and the
piscivorous fish F. commersonii.
The third major cluster (blue colour in Figure 17) included alien molluscs
such as B. pharaonis, Chama pacifica, and Magallana/Crassostrea species,
the alien polychaetes F. enigmaticus and H. elegans, the alien macroalgae
Asparagopsis spp. and U. pinnatifida, and the cryptogenic hexacoral Oculina
patagonica. A high number of impacts on biodiversity was recorded for
these species, mainly through the mechanisms of competition and the
creation of novel habitat (both negative and positive for the latter) as
structural ecosystem engineers.
The final cluster (light blue colour) was delineated by 45 species with
394 impact records. Thirty-six species of this cluster had less than 10
records each. Competition for resources was the most common mechanism
of impact with 59 records, caused by 33 of the 46 species of the cluster.
Fouling species with negative impacts on aquaculture followed as the second
most common mechanism. Such species with both competition and fouling
impacts were the ascidians Styela plicata, Clavelina oblonga, and Didemnum
vexillum, the bivalve P. radiata, the sponge P. magna, and the bryozoan
T. inopinata. Most species with negative impacts on human health were
also included in this cluster. Diadema setosum, M. philippina, P. lineatus, and
P. miles were grouped together with high similarity as stinging species.
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Figure 18 . Type of evidence of the analysed impact records on biodiversity, ecosystem
services and human health. The numbers on top of the bars indicate the sample size (number of
reported cases) for each category.
Type of evidence
Only 16% of impact records were experimental (9% were based on
manipulative and 7% on natural experiments; Figure 18). Moreover, many
of the experimental studies were in vitro experiments on alien species
extracts and compounds. Most of the taxa studied by experimental studies
were Chlorophyta and Rhodophyta. Overall, direct observation was the
most common type of evidence accounting for 40% of all records. Reported
impacts on provisioning ecosystem services were almost exclusively based
on direct observations (Figure 18). Fouling on aquaculture facilities,
entanglement in nets, and new commodities in the fisheries and fish
markets are undoubtedly direct impacts and as such were classified as
direct observations. For biodiversity records, direct observations referred
to the creation of novel habitat, apparent competitive interactions such as
overgrowth of sessile organisms, predation impacts derived from stomach
content analysis, and massive mortalities. Expert judgement and non-
experimentally based correlations accounted for 32% and 9% of impact
records, respectively (Figure 18), the former being dominant for “regulating
and maintenance” ecosystem services. Modelling overall accounted for
only 2% of the records.
The western Mediterranean and the Alboran Sea exhibited the highest
percentages of experimental studies assessing negative impacts on biodiversity,
Bioinvasions impacts in the Mediterranean Sea
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Figure 19 . Geographical variation of type of evidence for negative (A) and (B) positive impacts
on biodiversity across Mediterranean ecoregions.
followed by the Adriatic and the Ionian Sea (Figure 19A). Much less
experimental studies were recorded in the Aegean and the Levantine Sea.
Instead, expert judgement was the main type of evidence in these
ecoregions, as well as in the Adriatic and the Ionian Sea. The western
Mediterranean and the Adriatic Sea were the ecoregions with the highest
proportion of experimental studies for positive impacts on biodiversity
(Figure 19B). Most impact reporting through modelling studies for both
negative and positive impacts was from the Levantine Sea (Figure 19A, B).
Inventory of high impact species in the Mediterranean Sea
An inventory of impactful species, based on both negative and positive
reported impacts on biodiversity, ecosystem services and human health, is
proposed in Table 4 and includes 88 species. Of the 103 assessed species, 15
were excluded from this inventory due to the complete absence of impact
records or the reporting of only minor impacts. For three included species
there was only weak evidence for their impact, based on expert judgement.
Based on their negative impacts score on biodiversity, a list of the ten worst
invasive species was compiled (Table 5). Caulerpa cylindracea ranked first,
followed by W. setacea and L. lallemandii. Macroalgae dominated in the
list of the ten worst invasives (6 species), followed by fishes and bivalve
molluscs (2 species each).
Discussion
The number of reported negative impact records on biodiversity was
approximately four times higher than the number of positive ones.
Analysing the mechanisms of these effects, it was revealed that many alien
species strongly compete for space and resources with native biota (e.g.,
Turon et al. 2007; Fanelli et al. 2015; Manconi et al. 2020) whereas others
act as ecosystem engineers and alter the structure of habitats causing major
negative effects on the local communities (e.g., Rilov et al. 2004; Bedini et
al. 2014). Furthermore, intense feeding activities of invasive species can
cause the decline of native populations (e.g., Kampouris et al. 2019), with
alien grazers even being able to cause massive negative impacts in eastern
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Table 4. Proposed inventory of alien and cryptogenic marine species with reported moderate to high impacts on biodiversity or
ecosystem services or human health. Species whose impact was only reported by expert judgement are marked in light yellow
(weak evidence); Cr: For cryptogenic species.
Cercozoa
Tracheohyta
Annelida: Polychaeta
Chordata: Osteichthyes
Haplosporidium pinnae (Cr)
Halophila stipulacea
Ficopomatus enigmaticus
Alepes djedaba
Foraminifera
Porifera
Hydroides elegans
Atherinomorus forskalii
Amphistegina lobifera
Paraleucilla magna
Arthropoda: Crustacea
Decapterus russelli
Ochrophyta
Cnidaria
Acartia (Acanthacartia) tonsa
(Cr)
Dussumieria elopsoides
Chrysonephos lewisii
Macrorhynchia philippina
Callinectes sapidus
Etrumeus golanii
Rugulopteryx okamurae
Oculina patagonica (Cr)
Dyspanopeus sayi
Fistularia commersonii
Sargassum muticum
Rhopilema nomadica
Erugosquilla massavensis
Herklotsichthys punctatus
Stypopodium schimperi
Ctenophora
Metapenaeus monoceros
Lagocephalus sceleratus
Undaria pinnatifida
Mnemiopsis leidyi
Metapenaeus stebbingi
Nemipterus randalli
Chlorophyta
Bryozoa
Paracerceis sculpta
Parupeneus forsskali
Caulerpa cylindracea
Amathia verticillata (Cr)
Penaeus aztecus
Pempheris rhomboidea
Caulerpa taxifolia
Tricellaria inopinata
Penaeus pulchricaudatus
Plotosus lineatus
Caulerpa taxifolia var. distichophylla
Mollusca
Penaeus semisulcatus
Pterois miles
Cladophora patentiramea
Anadara kagoshimensis
Portunus segnis
Sargocentron rubrum
Codium arabicum
Anadara transversa
Echinodermata
Saurida lessepsianus
Codium fragile subsp. fragile
Arcuatula senhousia
Diadema setosum
Scomberomorus commerson
Codium parvulum
Brachidontes pharaonis
Chordata: Ascidiacea
Siganus luridus
Halimeda incrassata
Bursatella leachii (Cr)
Ciona robusta
Siganus rivulatus
Rhodophyta
Chama pacifica
Clavelina oblonga
Sphyraena chrysotaenia
Acrothamnion preissii
Conomurex persicus
Didemnum vexillum
Torquigener flavimaculosus
Agarophyton vermiculophyllum
Fulvia fragilis
Herdmania momus
Upeneus moluccensis
Antithamnion nipponicum
Magallana/Crassostrea species
Microcosmus squamiger
Upeneus pori
Asparagopsis armata
Pinctada radiata
Polyandrocarpa zorritensis
Asparagopsis taxiformis
Rapana venosa
Styela plicata
Galaxaura rugosa
Ruditapes philippinarum
Grateloupia turuturu
Spondylus spinosus
Lophocladia lallemandii
Womersleyella setacea
Table 5. The ten worst invasive species, based on their negative impact
score (accounting only for impacts on biodiversity). This ranking does not
account for the spatial extent of impacts but is based on a sum of all
reported impacts in the literature, weighted by their magnitudes.
Species
Negative Impacts Score
Caulerpa cylindracea
134
Womersleyella setacea
80
Lophocladia lallemandii
65
Brachidontes pharaonis
47
Siganus luridus
44
Rugulopteryx okamurae
43
Caulerpa taxifolia
38
Siganus rivulatus
38
Acrothamnion preissii
36
Spondylus spinosus
33
Mediterranean ecoregions by depleting rocky reef algal biomass (e.g., Sala
et al. 2011). At the same time, some alien species create novel habitats that
offer shelter to many other native organisms (e.g., Di Martino et al. 2007;
Munari et al. 2015), constitute new food sources for native predators (e.g.,
Giacoletti et al. 2016; Tiralongo et al. 2021), and contribute to other
ecological functions of the recipient ecosystem (e.g., Sarà et al. 2021).
It is noteworthy that a higher number of positive than negative impacts
were recorded affecting the flow of ecosystem services. Food provision
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appears as the most affected ecosystem service by both positive and
negative impacts. Invasive fish have significantly increased their numbers
in the last thirty years, representing today more than half of the total
abundance and biomass of the Levantine fish (Katsanevakis et al. 2018).
Simultaneously, severe population declines in some native fish populations
occurred (Edelist et al. 2013). All surveys that indicated a competitive effect
of alien species on commercial native species stocks were supported by
weak or modest strength of evidence. There is no strong evidence supporting
the hypothesis of native species population declines exclusively or largely
due to competition with alien species. On the other hand, research has
demonstrated that many species have declined in the Levantine Sea due to
climate change, as the new temperature regime is beyond their thermal
niche (Rilov 2016). Climate change is an important stressor for native biota
and has negatively affected the Mediterranean Sea (Marbà et al. 2015; Rilov
2016; Albano et al. 2021). In the western Mediterranean, shallow waters
have been warming for more than a century with more abrupt positive
trends during the recent decades that have resulted in local sea surface
temperature increases that even exceed 1 °C (Lejeusne et al. 2010) and
deleterious marine heatwaves (Garrabou et al. 2009, 2019, 2022). The
eastern basin was shown to be warming faster (Pisano et al. 2020; Novi et
al. 2021), with surface water temperature even exceeding a 3 °C increase
during the period 1978–2014 in the southeastern regions (Ozer et al. 2017).
Climate change has played a crucial role in the alteration of eastern
Mediterranean ecosystems and the collapse of several species (Rilov 2016;
Albano et al. 2021), but these dramatic changes are better explained as a
result of the combined effects of climate change and biological invasions
(Marras et al. 2015; Azzurro et al. 2019; Yeruham et al. 2020). Indeed,
multiple stressors often act in synergy, leading to multiplicative effects
(Korpinen et al. 2019; Gissi et al. 2021).
A progressive transition towards a more thermophilic biota is reported
for the global fishery catch for most marine ecosystems (Cheung et al.
2013), and also in the Mediterranean Sea (Tsikliras and Stergiou 2014),
where warm adapted invaders of Indo-Pacific origin are already replacing
native biota (Stergiou et al. 2016 and references therein; Katsanevakis et al.
2018; Albano et al. 2021). Yet, to date many alien species contribute to food
provision as new commodities (Katsanevakis et al. 2018), especially in the
eastern Mediterranean where alien fishes and crustaceans are more
abundant in comparison to the western basin (Katsanevakis et al. 2014b).
Levantine fisheries have even begun to shift their target from deeper to
shallower waters where they can catch larger quantities of thermophilic
alien species (van Rijn et al. 2020).
Contrary to the general idea that invading alien species are inherently
“bad”, a view that is often dependent on social perceptions (Katsanevakis et
al. 2014a), the positive effects of tropical invaders are worth exploring in
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detail, especially in the current climate change context. In the Levantine
Sea, the ecological space made available by the shift of temperate species
towards higher latitudes and greater depths cannot be covered by
neonatives (sensu Essl et al. 2019) from southern regions, as this is a land-
locked basin. In that sense, it can be argued that thermophilic Lessepsian
immigrants entering the Mediterranean through the Suez Canal may
contribute to limiting the loss of ecosystem functions and services. Restoring
the Levantine Sea ecosystems to their previous state before climate change and
massive biological invasions is unrealistic. Except for targeted management
actions for specific species (e.g. P. miles and L. sceleratus), marine
managers in the region need to be adaptive and realistic, understanding
and accepting the role of alien species in the novel marine ecosystems.
Regarding the impacts of invasive species on human health, we observed
that these are mostly limited to the eastern basin. Among the nine species,
only L. sceleratus was found to have an impact in locations west of the
Aegean Sea, and the species has already reached the westernmost end of
the Mediterranean (Azzurro et al. 2020). The actual number of negative
impact cases on human health is probably underestimated because many
non-life-threatening episodes do not require hospitalisation and are
treated by first aid. Such cases are most often not officially reported (Bédry
et al. 2021), especially not by scientific literature. Bédry et al. (2021)
proposed a series of actions to manage the effects of marine biological
invasions on human health: i) to reinforce public awareness of the potential
dangers that specific alien species may pose to human health, ii) to train
and prepare professional medical personnel on the physiological effects of
toxic alien species, iii) to establish a regional warning and monitoring
system. In this review, we did not record positive impacts of alien or
cryptogenic species on human health but future research on alien species is
expected to disclose their potential as providers of biomolecules, to be used
by the pharmaceutical industry (e.g. Genovese et al. 2009; Minicante et al.
2016; Dhahri et al. 2020). However, the discovery of new potentially active
molecules is a well-worn argument very often put forward to minimise the
problem of species introductions (a “chestnut” in journalistic terms), but
its validity has only rarely, if ever, been demonstrated and evaluated.
Marine species capable of producing active metabolites are well known
around the world and invasions are not necessary to initiate research and
exploitation. The industrial exploitation of such molecules often requires
either aquaculture production or chemical synthesis. So far, to the best of
our knowledge, there are no examples of industrial exploitation of
Mediterranean invaders for active molecules.
Biological invasions remain one of the most catastrophic threats to
global biodiversity, but the complexity of their impacts along with their
ecological, and socio-economic dimensions are worth exploring in detail
and at different scales. Some scientists consider that the origin of a species
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does not make it de facto responsible for negative consequences (Davis et
al. 2011) since native species also possess the ecological ability to affect
ecosystems and cause deleterious impacts under certain conditions. For
example, the coastal rocky reefs of the eastern Mediterranean are
overgrazed by the invasive rabbitfishes Siganus luridus and S. rivulatus
(Sala et al. 2011; Giakoumi 2014; Vergés et al. 2014; Yeruham et al. 2020)
but under certain conditions, similar effects can be provoked by native
grazers (Hereu 2006; Vergés et al. 2009; Bonaviri et al. 2011; Gianni et al.
2017; Tsirintanis et al. 2018; Papadakis et al. 2021), often due to cascading
effects caused by the loss of overfished native predators (Sala et al. 1998). It
should also be highlighted that some of the high-impact species of the
current proposed inventory are cryptogenic, meaning that their origin
cannot be completely ascertained. Non-factual perception of alien species
as invasive (i.e. negatively impacting ecosystems and human well-being)
neglects a holistic approach to impact assessment and can lead to bias
(Goodenough 2010; Katsanevakis et al. 2014a). Commonly the same species
have both negative and positive impacts on biodiversity, and the overall
balance is hard to assess. For example, the possible facilitation of the
restoration of a degraded ecosystem by an invasive alien species like the
facilitation of seagrass seedlings by C. cylindracea in degraded meadows
(Ceccherelli and Campo 2002; Pereda-Briones et al. 2018) deserves further
consideration and confirmation by long-term studies evaluating the survival
rate of germlings in dense C. cylindracea meadows. This brings up ecological,
cultural, and socio-economic considerations that complicate decision
making in invasive species management, as changes in host communities
are not always perceived as harmful (Bonanno 2016), and management
decisions need to account for conflicts and trade-offs (Mason et al. 2017).
In this study, we found a clear geographic differentiation in the distribution
of reported impacts. For example, predation as a mechanism of impact was
more important in the Levantine Sea than in the rest of the basin. This is
following the higher richness and abundance of invasive fishes that
characterise the eastern sectors of the Mediterranean, being introduced
from the Red Sea through the Suez Canal and favoured by the higher
temperatures of the eastern Mediterranean (Katsanevakis et al. 2014b). The
great taxonomic differences of alien species among ecoregions (Zenetos et
al. 2012; Katsanevakis et al. 2014b), as well as the significant differences in
impacts among taxonomic groups, revealed in this study, partly explain the
variation detected among Mediterranean ecoregions.
A global effect of alien species invasions is biotic homogenization, the
process of native biodiversity impoverishment happening simultaneously
with a proliferation of alien species that results in the expansion of
homogeneity in terms of genetic, taxonomic and functional diversity with
unclear biodiversity distinctions (Olden et al. 2004 and references therein).
Signs of biotic homogenization are evident in the marine ecosystems of the
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Mediterranean Sea, as the establishment and advancement of biological
invasions in the basin have occurred alongside native biota impoverishment
and losses (Galil 2007 and references therein). In many cases, an impoverished
homogeneous state has been spread within the Mediterranean basin, through
the spread and dominance of some alien structural ecosystem engineers
(Navarro-Barranco et al. 2018; Morri et al. 2019). The increasing rate of
biological invasions in the Mediterranean and the continuing seawater
warming led to the replacement of endemic populations by mostly
thermophilic Lessepsian species. Hence, the “Mediterranean character” of
the marine biota progressively declines, contributing to the breakdown of
the regional distinctiveness of the Mediterranean Sea (Ben Rais Lasram and
Mouillot 2009).
In the current literature review, most studies on impacts on biodiversity
were based on medium and low strength of evidence. In agreement with
Katsanevakis et al. (2014a), the proportion of reported impacts with high
strength of evidence (i.e. manipulative and natural experiments) was low.
These studies were conducted in 13 out of the 22 Mediterranean countries,
with most cases (84%) reported from Italy, Spain, and France (Figure 20A).
The latter are the Mediterranean countries with the higher budget for
research and development (UNESCO 2021) (Figure 20B), reflecting the
connection between research robustness, number of experimental studies,
and investment in research and development. Indeed, several studies relate
scientific output to national spending on research and development, the
number of universities in a country, GDP, and English proficiency (Man et
al. 2004; Meo et al. 2013; Jamjoom and Jamjoom 2016; Mueller 2016).
Another possible reason for the higher inferential strength of impact
assessment studies in the western Mediterranean countries could be related
to the taxonomic distribution of alien species within the Mediterranean.
There is a higher species richness of alien macroalgae in the western
Mediterranean in comparison to the eastern basin, and the opposite
pattern for alien fish species, which have primarily invaded the eastern
Mediterranean (Katsanevakis et al. 2014b). Our analysis revealed that the
vast majority (90%, n = 76) of experimental impact cases on biodiversity
reported from the western Mediterranean and the Alboran Sea assess
impacts caused by alien macroalgae, whereas no experimental study
investigates impacts caused by alien fish. At the same time, experimental
impact records are significantly lower in numbers in the Levantine and the
Aegean Sea and mainly refer to alien fish effects (71%; n = 14). Macroalgae
are sessile organisms, easier to be controlled in confined experimental
treatments, in contrast to mobile organisms such as fish. There is already a
high level of difficulty in performing underwater experiments to investigate
ecological interactions due to the various factors that affect a diver’s
awareness such as visibility and temperature (O’Brien and Caramanna
2017). In addition, species mobility may interfere with the success of an
Bioinvasions impacts in the Mediterranean Sea
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Figure 20. A) Number of reported experimental studies included in this review assessing
impacts on biodiversity, per Mediterranean country, B) Total intramural expenditure on Research
and Development (R&D), performed by some Mediterranean countries during 2018, expressed
in Purchasing Power Parity dollars (UNESCO 2021).
applied underwater experiment and this fact can explain why experimental
research on alien species is more limited in the eastern Mediterranean,
dominated by mobile species invasions. Similar results were reported by
Kytinou et al. (2020), who reviewed methodological approaches globally
that analyse coastal shelf food webs, and concluded that only 14% of the
reviewed studies conducted manipulative or natural experiments, and most
of the studies focused on benthic species and much less on fish. Also,
Thomsen et al. (2014) reviewed aquatic alien species’ experimental impact
assessment surveys within forty years (1972–2012) and found only one
alien marine fish species studied by experiments among more than a
hundred different species. The species mobility hypothesis is further
supported by the fact that in the Levantine and the Aegean Sea most of the
impacts were reported by experimental studies on the herbivorous fishes
Siganus spp. with assessments based on the effects on macroalgal
communities, i.e. the sessile food of siganids.
Bioinvasions impacts in the Mediterranean Sea
Tsirintanis et al. (2022), Aquatic Invasions 17(3): 308–352, https://doi.org/10.3391/ai.2022.17.3.01 342
Furthermore, there are several additional confounding factors that
distort the actual impact patterns, such as paucity of data (Ojaveer et al.
2015), taxonomic impediment (Engel et al. 2021), challenges for assessing
and quantifying the actual impact of particular alien taxa (Carlton 2002;
Simberloff 2011; Davidson and Hewitt 2014; Strayer et al. 2017), and weak
inferential strength of evidence (Katsanevakis et al. 2014a). Another aspect
to consider is that the current analysis examines impacts caused by each
alien species separately. Βut species do interact, and their cumulative
impacts are not necessarily additive but can be multiplicative or mitigative
(Katsanevakis et al. 2016), also often interacting with other local or global
stressors such as climate change (Gissi et al. 2021). Hence, impacts by alien
species reported in a specific location are not necessarily representative of
the entire ecoregion. As a consequence, spatial patterns herein reported
should be interpreted with caution, as highlighted in other studies reporting
spatial patterns in biological invasions (e.g. Pyšek et al. 2008; Stranga and
Katsanevakis 2021).
In addition, the methodological approach of the current impact
assessment framework is performed without chronological restrictions and
depicts the effects of alien and cryptogenic species on biodiversity,
ecosystem services and human health from the moment of the first impact
report of each targeted species until 2021. The temporal variation of
impacts was not taken into account. Still, invasive alien species can regress
after an invasion phase, following a so-called “boom and bust” invasion
pattern, as has been reported for Caulerpa cylindracea, C. taxifolia,
L. lallemandii and S. schimperi (Iveša et al. 2006; Montefalcone et al. 2015;
Dimitriadis et al. 2021; Santamaría 2021). Unfortunately, the literature of
marine invasions sorely lacks scientific articles showing the evolution of
invaded sites over several decades (Strayer et al. 2017; Ojaveer et al. 2018).
The reason may be due to the functioning of scientific research, i.e. the
short duration of doctoral thesis projects, short-term funding, the need to
publish quickly, the lack of interest of scientific journals and reviewers in
what seems routine monitoring, and probably also the fact that spectacular
bad news is easier to publish than good news. Hence, it is possible that
some of the published impacts later declined in magnitude, due to the
“boom-and-bust” dynamics of biological invasions. This stresses the need for
repetitive assessments and regular re-evaluation of the impacts of alien species.
We have provided a list of the “ten worst invasive species” in the
Mediterranean, based on reported impacts. Similar lists have been
compiled in the past (e.g. see Streftaris and Zenetos 2006; Katsanevakis et
al. 2016) based on different approaches and metrics. Depending on the
criteria applied for impact assessment, different schemes can lead to
strikingly different outcomes (González-Moreno et al. 2019). Hence, any
such ranking of invasive species should be perceived under the limitations
and assumptions of the analysis that produced it. For example, the present
Bioinvasions impacts in the Mediterranean Sea
Tsirintanis et al. (2022), Aquatic Invasions 17(3): 308–352, https://doi.org/10.3391/ai.2022.17.3.01 343
study did not account for the spatial distribution of impactful species and
the geographical extent of reported impacts, which for some species can be
quite small (i.e. high impact but in a restricted area). Nevertheless, it is
remarkable that our inventory of the ten worst invasive species is quite
similar to the respective ranking in Katsanevakis et al. (2016) [their
indicator D3], produced by a quite different approach, i.e. the sum of
cumulative impact scores of the species on marine habitats based on a
conservative additive model (CIMPAL) across all 10 × 10 km cells that
cover the entire Mediterranean Sea. Eight of the ten top species are shared
by the two rankings.
We recorded a variety of impacts on biodiversity, ecosystem services and
human health caused by alien and cryptogenic species. Without forgetting
that all introduced species represent a disturbance to nature by modifying
native ecosystems (Ricciardi et al. 2013), management choices need to be
based on robust evidence. This remains a primary need in the process of
evaluating the effects of alien species in ecosystems. The way forward to
interpret biological invasions and prioritise management actions is to
incorporate holistic approaches to alien species impact assessments,
accounting for both positive and negative impacts and socio-ecological
tradeoffs, and increase the strength of evidence through the implementation
of more experimental studies.
Acknowledgements
We thank N. Koukourouvli for her assistance in GIS and Michèle Perret-Boudouresque for
documentation assistance. We thank the two anonymous reviewers for their comments and
suggestions that significantly contributed to the improvement of the manuscript.
Funding declaration
The present study was supported by the Hellenic Foundation for Research and Innovation
(H.F.R.I.) under the “First Call for H.F.R.I. Research Projects to support Faculty members and
Researchers and the procurement of high-cost research equipment grant” (Project ALAS –
“ALiens in the Aegean – a Sea under siege”; Project Number: HFRI-FM17-1597; Katsanevakis
et al. 2020b). FaCr was partially funded by the project PO FEAMP CAMPANIA 2014-2020
(DRD n.35 of 15th March 2018). XT obtained partial funding from project MARGECH
(PID2020-118550RB, MCIN/AEI/10.13039/501100011033) from the Spanish Government. AR
received grants from the University of Catania through “PiaCeRi-Piano Incentivi per la Ricerca
di Ateneo 2020–22 linea di intervento 2”.
Authors’ contribution
KTs: research conceptualization, design and methodology, review and data collection, data
analysis, writing the first draft of the manuscript; SK: research conceptualization, design and
methodology, data analysis, supervision, editing and writing; all authors: validating data,
contribution with additional data and taxonomic expertise, interpretation, editing and writing.
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Supplementary material
The following supplementary material is available for this article:
Appendix 1. Species-specific review of the impacts of alien and cryptogenic species on biodiversity, ecosystem services and human
health in the Mediterranean Sea, and other related information.
This material is available as part of online article from:
http://www.reabic.net/aquaticinvasions/2022/Supplements/AI_2022_Tsirintanis_etal_Appendix.pdf
1
Bioinvasion impacts on biodiversity, ecosystem services, and human health in
the Mediterranean Sea
Konstantinos Tsirintanis1, Ernesto Azzurro2,3, Fabio Crocetta3, Margarita Dimiza4, Carlo Froglia2, Vasilis
Gerovasileiou5,6, Joachim Langeneck7, Giorgio Mancinelli8,9,10, Antonietta Rosso10,11, Nir Stern12, Maria
Triantaphyllou4, Konstantinos Tsiamis13, Xavier Turon14, Marc Verlaque15, Argyro Zenetos16 and Stelios
Katsanevakis1,*
1Department of Marine Sciences, University of the Aegean, Lofos Panepistimiou, 81100 Mytilene, Greece
2CNR-IRBIM, National Research Council. Institute of Biological Resources and Marine Biotechnologies, Ancona,
Italy
3Department of Integrative Marine Ecology, Stazione Zoologica Anton Dohrn, I-80121 Naples, Italy
4Faculty of Geology & Geoenvironment, National and Kapodistrian University of Athens, Panepistimioupolis 15784,
Athens, Greece
5Department of Environment, Faculty of Environment, Ionian University, 29100 Zakynthos, Greece
6Hellenic Centre for Marine Research (HCMR), Institute of Marine Biology, Biotechnology and Aquaculture
(IMBBC), 71500 Heraklion, Greece
7Department of Biology, University of Pisa, 56126 Pisa, Italy
8Department of Biological and Environmental Sciences and Technologies (DiSTeBA), University of Salento, SP
Lecce-Monteroni, 73100 Lecce, Italy
9National Research Council – Institute of Marine Biological Resources and Biotechnologies (CNR-IRBIM), 71010
Lesina (FG), Italy
10CoNISMa, Consorzio Nazionale Interuniversitario per le Scienze del Mare, 00196 Roma, Italy
11Department of Biological, Geological and Environmental Sciences, University of Catania, 95129 Catania, Italy
12Israel Oceanographic and Limnological Research, National Institute of Oceanography, Haifa 31080, Israel
13Vroutou 12 Athens, 11141, Greece
14Centre for Advanced Studies of Blanes (CEAB, CSIC), 17300 Blanes, Catalonia, Spain
15Aix Marseille University and Université de Toulon, CNRS, IRD, Mediterranean Institute of Oceanography (MIO),
UM 110, Marseille, France
16Institute of Marine Biological Resources & Inland Waters (IMBRIW), Hellenic Centre for Marine Research
(HCMR), 16452, Argyroupolis, Greece
*Corresponding author, e-mail: stelios@katsanevakis.com
Supplementary material. Appendix 1. Species-specific review of the impacts of alien and
cryptogenic species on biodiversity, ecosystem services and human health in the Mediterranean
Sea, and other related information.
Recommended citation:
Tsirintanis K, Azzurro E, Crocetta F, Dimiza M, Froglia C, Gerovasileiou V, Langeneck J, Mancinelli G, Rosso A, Stern
N, Triantaphyllou M, Tsiamis K, Turon X, Verlaque M, Zenetos A, Katsanevakis S (2022) Bioinvasion impacts on
biodiversity, ecosystem services, and human health in the Mediterranean Sea. Aquatic Invasions 17(3): 308–352,
https://doi.org/10.3391/ai.2022.17.3.01
2
Contents
Cercozoa........................................................................................................................................................3
Foraminifera..................................................................................................................................................4
Ochrophyta....................................................................................................................................................4
Chlorophyta...................................................................................................................................................8
Rhodophyta.................................................................................................................................................13
Tracheophyta..............................................................................................................................................21
Porifera........................................................................................................................................................22
Cnidaria.......................................................................................................................................................23
Ctenophora.................................................................................................................................................24
Bryozoa........................................................................................................................................................25
Mollusca......................................................................................................................................................26
Annelida:Polychaeta...................................................................................................................................37
Arthropoda:Crustacea................................................................................................................................38
Echinodermata............................................................................................................................................43
Chordata:Ascidiacea...................................................................................................................................44
Chordata:Osteichthyes...............................................................................................................................46
References...................................................................................................................................................54
3
Cercozoa
Haplosporidium pinnae Catanese et al. 2018
Haplosporidium pinnae is a protozoan of unknown origin, firstly reported from Spain in the
western Mediterranean in 2016, infecting tissues of the Mediterranean endemic bivalve Pinna
nobilis Linnaeus, 1758 (Darriba 2017; Vázquez-Luis et al. 2017; Catanese et al. 2018). Later,
Scarpa et al. (2020) reported the presence of the protozoan to individuals of the bivalves Ruditapes
decussatus (Linnaeus, 1758) and Mytilus galloprovincialis Lamarck, 1819 from Sardinia. These
R. decussatus specimens constitute the oldest evidence of H. pinnae presence in the Mediterranean,
having been sampled in 2014 (Scarpa et al. 2020). In autumn 2016, a large-scale mass mortality
event (MME) of Pinna nobilis rapidly spread throughout the Spanish Mediterranean coasts and by
the summer of 2017, fan mussel populations had completely collapsed in the southern coastal areas
(Vázquez-Luis et al. 2017). H. pinnae was reported to systemically infect individuals of the local
population (Darriba 2017) and was associated with the catastrophic mortality event (Vázquez-Luis
et al. 2017; Catanese et al. 2018). Evidence suggest that the parasite was dispersed to the western
Mediterranean through surface currents (Cabanellas-Reboredo et al. 2019). By 2018 the MMEs
had progressed to the Italian coasts up until the Ionian Sea at the gulf of Taranto (Catanese et al.
2018; Panarese et al. 2018; Carella et al. 2019; Tiscar et al. 2019; Betti et al. 2021). However,
Carella et al. (2019) reported systemic inflammatory lesions in the fan mussel tissues that were
linked to a Mycobacterium sp. as the causative agent of mortalities in Campania and Sicily. During
the same year (2018), MMEs were also reported from the northern Aegean Sea, and H. pinnae was
detected to parasitize and infect local fan mussel populations (Katsanevakis et al. 2019). Lattos et
al. (2020) reported the simultaneous detection of H. pinnae and a Mycobacterium sp. from samples
of this region. In 2019, the MMEs reached the Adriatic Sea, where both H. pinnae and
Mycobacterium sp. were traced in sampled fan mussels and were identified to infect P. nobilis
specimens (Čižmek et al. 2020; Šarić et al. 2020). More recently, some Vibrio bacteria have also
been associated with captivity or winter mortalities of the fan mussels (Prado et al. 2020a; Lattos
et al. 2021). It is possible however that multiple microbes are involved in the pathogenesis of the
lethal pandemic episodes throughout the Mediterranean (Carella et al. 2020; Lattos et al. 2020;
Šarić et al. 2020; Scarpa et al. 2020). Furthermore, Box et al. (2020) analyzed histologically and
molecularly P. nobilis specimens from the western Mediterranean, sampled before the disease
outbreak (2011-2012) and after the outbreak (2016-2017). H. pinnae infected all specimens
collected after the mortality outbreak and was absent in the tissues of the earlier collected samples.
Box et al. (2020) showed that apart from heavily infecting the digestive system of P. nobilis, H.
pinnae is responsible for a collapse of the antioxidant defenses of the animals that favours
instauration of oxidative stress and cellular damage. High mortality rates could be the effect of
such an infection by H. pinnae and could prove more harmful in the case of a multiple co-infection
with other pathogens such as Mycobacterium sp. (Box et al. 2020). The results of the MMEs
throughout the Mediterranean have proved devastating and P. nobilis faces extinction from the
coastal ecosystems that it occupied before 2016 (Katsanevakis et al. 2021). Some natural refugia
still exist at a few locations of the Mediterranean with specific environmental conditions, but even
there fan mussel numbers continue to decline (Zotou et al. 2020; Katsanevakis et al. 2021). Due to
4
the sharp decline of its populations and the threat that the species faces with a high risk of
extinction, P. nobilis was recently classified as Critically Endangered (Kersting et al. 2019).
Foraminifera
Amphistegina lobifera Larsen, 1976
The Lessepsian foraminifer Amphistegina lobifera has rapidly established invasive populations
throughout the eastern Mediterranean basin during the last decades and is currently thriving on the
doorstep of the western Mediterranean (Triantaphyllou et al. 2009; Mouanga 2018; Guastella et
al. 2019; Stulpinaite et al. 2020). A. lobifera lives predominantly on hard and phytal substrates in
the shallow environments and prefers mid to high light intensity for reproduction (Hallock 1981;
Triantaphyllou et al. 2012). Its distributional range and settlementis strongly related to water
temperature and is generally delimited by the 13.7 oC winter temperature isotherm (Hollaus and
Hottinger 1997; Langer and Hottinger 2000; Triantaphyllou et al. 2012). Under the rates of current
climate change this invasive species is expected to invade the western Mediterranean and the
Adriatic Sea during the 21st century and may trigger significant changes on the native foraminiferal
assemblages and ecosystem functioning (Weinmann et al. 2013). On its current expansion range,
A. lobifera has affected native epifaunal communities by establishing abundant populations that
dominate and alter community structure (Siokou et al. 2013; Caruso and Cosentino 2014; Mouanga
2018). At some occasions, A. lobifera has formed such densities that it completely alters ecosystem
structure by changing the substrate to a monospecific layer of its shells (Streftaris and Zenetos
2006; Yokeş et al. 2014; Vohník 2021). The creation of such novel habitats contributes to sediment
stability (Triantaphyllou et al. 2009, 2012; Mouanga 2018) and furthermore, as carbonate
producer, large aggregations of A. lobifera significantly contribute to carbon sequestration.
Amphisteginid foraminifers have been assessed as indicator species due to sensitivity to marine
pollution and as such, several studies in the Mediterranean have correlated A. lobifera presence or
shell morphological deformities with eutrophication or heavy metal pollution (Yanko et al. 1998;
Al-Salameen 2001; Koukousioura et al. 2011; Meriç et al. 2012; El Kateb et al. 2018).
Ochrophyta
Chrysonephos lewisii (W.R.Taylor) W.R.Taylor
This filamentous chrysophyte of tropical Atlantic origin has been introduced into the
Mediterranean Sea, reported from Corsica and Italian coasts (Verlaque et al. 2015). The species
can give blooms during spring, contributing to the formation of extensive filamentous aggregates
covering seagrass meadows, macroalgal forests and gorgonian species, resulting in negative effects
to native communities (Sartoni et al. 1995; Hoffmann et al. 2000; Giuliani et al. 2005). These
aggregates also constitute a threat for the reef-building hexacoral Cladocora caespitosa (Linnaeus,
5
1767), as they reduce its colonies density, coverage and cause extensive necrosis (Kružić and
Požar-Domac 2007; De Biasi et al. 2021).
Rugulopteryx okamurae (E.Y.Dawson) I.K.Hwang, W.J.Lee & H.S.Kim
Rugulopteryx okamurae is a brown alga native in the North-western Pacific Ocean. It has been
recorded in France, Gibraltar, Morocco and Spain (Verlaque et al. 2009, 2015; Altamirano et al.
2016). The introduction pathway is probably the aquaculture of the Japanese oyster
Magallana/Crassostrea species. Since its first record in European waters in 2002 in Thau lagoon,
R. okamurae has been established there in great densities, mixed with many other NW Pacific
algae. In 2015, R. okamurae was first recorded in Mediterranean open waters on the south side of
the Strait of Gibraltar in the Sea of Alboran and rapidly exhibited invasive behaviour (Altamirano
et al. 2016; Ocaña et al. 2016; Sempere-Valverde et al. 2019; García-Gómez et al. 2020, 2021). It
can displace native species and fully homogenize the sea bottom, resulting in biodiversity loss and
change in the structure and composition of the native communities, including Cystoseira sensu
lato forests, Posidonia oceanica (Linnaeus) Delile meadows, coralligenous communities,
eulittoral and infralittoral communities of seaweeds, maërl communities, and epiphytic fauna of
invertebrates (Altamirano et al. 2017, 2019; Ocaña et al. 2016; El Aamri et al. 2018; García-Gómez
et al. 2018; CAGPYDS 2018, 2019; Ruitton et al. 2021). Several species of conservation interest
are also affected, such as the red coral Corallium rubrum (Linnaeus, 1758) and the endangered
species Patella ferruginea Gmelin, 1791 (García-Gómez et al. 2018; El Aamri et al. 2018). A few
years after its invasion in the Alboran Sea in 2015, R. okamurae dominated rocky coasts from very
shallow waters down to 50 m depth, replacing the native species Dictyota dichotoma (Hudson)
J.V. Lamouroux, deeply altering the composition of autochthonous algal assemblages, and
modifying the macrofaunal assemblages, contributing to a higher number of species and
abundance of individuals (Navarro-Barranco et al. 2019). When co-occurring with other invasive
macroalgae such as Caulerpa cylindracea Sonder, R. okamurae dominates and can restrict their
distribution (García-Gómez et al. 2018). The invasive macroalga is capable of causing community
shifts in the coralligenous habitats of Gibraltar (Sempere-Valverde et al. 2021). At Jbel Moussa,
R. okamurae became the dominant species and overlapped the native community in only one year,
leading to an increase in the number of dead colonies of the red gorgonian Paramuricea clavata
(Risso, 1826) and a significant decrease in the cover of the Corallinaceae Mesophyllum expansum
(Philippi) Cabioch & M.L.Mendoza (Sempere-Valverde et al. 2021). An important decline in the
conservation status of protected areas is expected where this seaweed settles. Economic impact of
the species is highly associated with losses in fisheries, tourism and beach management (MTERD
2020). Recreational and social services linked to touristic activities have been affected due to the
large amounts of biomass of the alga accumulated on beaches (Ruitton et al. 2021). Massive algal
drifts of R. okamurae have occurred in the city of Ceuta since 2017. In 2018, more than 5000
tonnes of R. okamurae biomass was removed, and fishers have reported the clogging of their nets
by the invasive macroalga that resulted in a decline in their catches (Sempere-Valverde et al. 2019).
No herbivory has been reported on this alga, which is strongly defended by anti-herbivory
diterpenes (Paula et al. 2011). R. okamurae contains compounds with antifungal, antibiotic,
antiinflammatory, insecticide and antiviral activity (Paula et al. 2011). The species has been
recently risk-assessed and proposed for possible inclusion in the list of Union concern of the EU
6
Regulation 1143/2014 on the prevention and management of invasive alien species (MTERD
2020).
Sargassum muticum (Yendo) Fensholt
First recorded in 1980 in the Mediterranean Sea (Thau Lagoon, France), Sargassum muticum is a
large perennial brown alga native to the northwestern Pacific. It is a successful invader worldwide
and was probably introduced in the Mediterranean as an epiphyte of Japanese oysters (Verlaque et
al. 2015). S. muticum has invaded transitional eutrophic environments and closed embayments of
the western Mediterranean and the Adriatic Sea. In coastal Lagoons (Thau, Venice) the species is
restricted to certain areas where it exhibits a seasonal development of caduc fronds in winter-spring
and impedes human activities, especially navigation (Boudouresque et al. 1985; Sfriso and Facca
2013; Sfriso et al. 2020). Its canopies are thick canopies and can reach several meters in height
(Curiel et al. 1998). S. muticum competes with indigenous species (Curiel et al. 1998; Occhipinti-
Ambrogi 2002). Furthermore, it is an important structural and light engineer that facilitates the
presence of some species, while decreasing the populations of others (Curiel et al. 1998;
Katsanevakis et al. 2014a; Smith et al. 2014; NEMESIS 2021 and references therein).In other
Atlantic regions, large S. muticum canopies introduced complexity to the recipient ecosystems and
were found to host a number of organisms and assemblages not necessarily constituting a
degradation in comparison to the previous state (Engelen et al. 2013; Belattmania et al. 2020;
Raoux et al. 2021). Nevertheless, the exceptionally long floating canopies collect both sediment
and floating debris, lowering aesthetic values (Sfriso and Facca 2013). Furthermore, S. muticum
fouls aquaculture structures and fishing gear, negatively impacting food provision (Verlaque 2001;
Boudouresque and Verlaque 2002; Katsanevakis et al. 2014a). Canopies of S. muticum even clog
intake pipes, hampering the operation of industrial facilities (Katsanevakis et al. 2014a; Bonanno
and Orlando-Bonaca 2019). The large thalli of the invasive perennial species may drift and wash
ashore, becoming a nuisance due to the foul smell of decomposition. Dense canopies may also
hinder swimming, boating and fishing. S. muticum has a fast growth rate combined with large size,
and stores a number of nutrients, thus negatively impacting ocean nourishment (Katsanevakis et
al. 2014a). On the other hand, it acts as a biofilter accumulating various harmful compounds,
especially in polluted and degraded ecosystems such as the Venice lagoon (Sfriso and Facca 2013).
Overall, there is a potential for various ecosystem services to be impacted (Katsanevakis et al.
2014a) such as food, biotic materials, coastal protection, cognitive benefits, recreation, symbolic
and aesthetic values, and life cycle maintenance (Salomidi et al. 2012). S. muticum has also
exhibited potential for commercial exploitation by the alginate industry (see ref. in Lewis 2009;
Josefsson and Jansson 2011). Extracted sulfated polysaccharides from S. muticum have shown a
great in vitro anti-leishmanial activity (Minicante et al. 2016). The alien ochrophyte also contains
antibacterial and antifungi substances (Plouguerné et al. 2008) fucoxanthins, antioxidants, and
other compounds of pharmaceutical interest (Milledge et al. 2016).
Stypopodium schimperi (Kützing) Verlaque & Boudouresque
Stypopodium schimperi is a Lessepsian brown alga that originates from the Red Sea, widespread
and abundant in the eastern Mediterranean basin. Its flat foliose formations occupy large surfaces
at rocky substrates and occasionally become invasive, monopolizing the habitat, and dominating
7
the macroalgal community (Boudouresque and Verlaque 2002; Tsiamis et al. 2010; Tsiamis 2012;
Verlaque et al. 2015; Bitar et al. 2017; Çinar et al. 2021). S. schimperi creates a novel habitat for
a variety of organisms, but at the same time it displaces previous space occupiers (Streftaris and
Zenetos 2006; Katsanevakis et al. 2014a; Smith et al. 2014). Having thin thalli and a high surface
to volume ratio, S. schimperi exhibits high nutrient uptake rates, resulting in a negative impact on
ocean nourishment and algal competitors (Katsanevakis et al. 2014a). Occasionally, large
quantities of S. schimperi wash ashore on coastal areas of the eastern Mediterranean negatively
affecting symbolic and aesthetic values and recreation and tourism (Streftaris and Zenetos 2006;
Israel et al. 2010; Hoffman et al. 2011; Katsanevakis et al. 2014a). The appearance and
proliferation of the invasive seaweeds Codium arabicum Kützing, Codium parvulum (Bory ex
Audouin) P.C.Silva, Galaxaura rugosa (J.Ellis & Solander) J.V.Lamouroux, and S. schimperi
alongside with marine pollution, have been associated with the replacement of native macroalgal
species (Hoffman et al. 2008, 2011). Phenolic substances, protein contents, carbohydrate yields
and fatty acid variations indicate that S. schimperii is a potential source for biological research,
new food resources, pharmacology and cosmetic industry (Sampli et al. 2000; Polat and Ozogul
2013; Ozgun and Turan 2015).
Undaria pinnatifida (Harvey) Suringar
The NW Pacific alga Undaria pinnatifida is an annual cold temperate heteromorphic brown alga
with a winter-spring development of macroscopic sporophytes. It was first recorded in the
Mediterranean Sea in 1971 (Thau Lagoon, France). U. pinnatifida was introduced, with oyster
imports from Japan, in transitional and eutrophic ecosystems of the western and central
Mediterranean as well as in the Adriatic Sea (Cecere et al. 2000; Verlaque et al. 2015). Undaria
pinnatifida causes multiple effects on the ecosystems it invades. It can decrease diversity through
competition with native species (Valentine and Johnson 2003, 2004; Hewitt et al. 2005; Farrell
and Fletcher 2006); in some instances the alien ochrophyte induces no changes to the recipient
communities, and in others acting as a structural ecosystem engineer, it can positively and/or
negatively affect a variety of species (Katsanevakis et al. 2014a; Smith et al. 2014). In the Mar
Piccolo di Taranto, the alien alga enhanced native biodiversity as a structural engineer in a polluted
and impoverished environment (Cecere et al. 2000). However, U. pinnatifida followed a boom and
bust invasion pattern in the lagoon with a rapid abundance increase during its first years of
introduction followed by a decline until local extinction (Cecere et al. 2016). In the Venice lagoon,
U. pinnatifida has been associated with a negative competitive impact for space with native
macroalgae and a decline of their numbers (Curiel et al. 1998, 2001; Occhipinti-Ambrogi 2002).
However, Sfriso and Faca (2013) found that U. pinnatifida and Sargassum muticum together
accounted for only 5–7 % of the total macroalgal biomass of the lagoon. The invasive alga also
fouls aquaculture facilities and fishing gear, inhibiting fish production and negatively affecting
food provision (Streftaris and Zenetos 2006; Katsanevakis et al. 2014a; Bonanno and Orlando-
Bonaca 2019). Its large, flat and comparatively thin thalli have rapid nutrient uptake, thus
negatively impacting ocean nourishment (Katsanevakis et al. 2014a). In polluted environments
such as the Venice lagoon, U. pinnatifida accumulates large amounts of various organic and
inorganic pollutants, contributing thus as a biofilter in water purification (Sfriso and Facca 2013).
Its large thalli have become a nuisance in the Venice canals, since they hinder navigation (Sfriso
8
and Facca 2013; Sfriso et al. 2020). U. pinnatifida polysaccharide extracts have been
acknowledged as a potential resource for anti-leishmanial and anticoagulant drugs for humans
(Faggio et al. 2015; Minicante et al. 2016). These compounds have also shown promising
antibacterial activity againstVibrio harveyi (Johnson and Shunk, 1936) that causes fish disease in
aquaculture (Rizzo et al. 2017a).
Chlorophyta
Caulerpa cylindracea Sonder
The SE Australian green macroalga Caulerpa cylindracea (previously accepted as Caulerpa
racemosa var. cylindracea (Sonder) Verlaque, Huisman & Boudouresque; Belton et al. 2014), was
first reported in the Mediterranean Sea in 1985 in Tunisia (Sghaier et al. 2016). It is now spread
throughout the Mediterranean Sea, becoming one of the most invasive species of the basin (Klein
and Verlaque 2008; Verlaque et al. 2015; Piazzi et al. 2016; Katsanevakis et al. 2016; Sghaier et
al. 2016; Zenetos et al. 2017). The invasive chlorophyte thrives in a wide range of depths and
environmental conditions, and in a variety of different habitats (Tsiamis 2012; Salomidi et al.
2012; Montefalcone et al. 2015a; Dailianis et al. 2016; Morri et al. 2019). At rocky reefs, C.
cylindracea mingles with turf and low-lying algae creating dense mats (100% cover) that
negatively affect indigenous communities through competition, overgrowth, increased
sedimentation, and ecosystem engineering (Piazzi et al. 2001, 2005, 2007;Piazzi and Cinelli 2003;
Balata et al. 2004; Baldacconi and Corriero 2009; Bulleri et al. 2010; Katsanevakis et al. 2010,
2014; Otero et al. 2013; García et al. 2015; Montefalcone et al. 2015b; Tamburello et al. 2015;
Betti et al. 2017; Cantasano et al. 2017; Longobardi et al. 2017; Mannino and Balistreri 2019;
Piazzi and Ceccherelli 2020). C. cylindracea novel habitats can substantially influence the biotic
assemblages of the invaded ecosystems, changing habitat structure and the structure of associated
faunal communities (Vázquez-Luis et al. 2009; Deudero et al. 2014). At the same time these new
complex turf structures create new habitats for many other organisms (Vázquez-Luis et al. 2009,
2013; Pacciardi et al. 2011; Katsanevakis et al. 2014a; Sinopoli et al. 2020). Rocky reef ecosystems
invaded by C. cylindracea in the western Mediterranean were estimated to support a wider trophic
niche in comparison to uninvaded ones, based on isotope signatures (Alomar et al. 2016). Sessile
organisms that lack effective defensive mechanisms, such as the sponge Sarcotragus spinosulus
Schmidt, 1862 and the coral Cladocora caespitosa, are particularly vulnerable to C. cylindracea
overgrowth (Kružić et al. 2008; Žuljević et al. 2011; Manconi et al. 2020). Even at low abundance,
the invasive chlorophyte competes for space with keystone species of the subtidal rocky reefs such
as the perennial canopy forming Cystoseira sensu lato (Sargassaceae, Ochrophyta) and while
pristine canopies may be resilient to C. cylindracea invasion, recovering perennial algae appear to
be more vulnerable (Bulleri et al. 2017). C. cylindracea, in association with Womersleyella setacea
(Hollenberg) R.E.Norris, negatively impacts the fitness of the gorgonian Paramuricea clavata
causing lower survival, higher necrosis rates and lower biomasses in invaded colonies. The
negative impact is enhanced in the case of P. clavata populations that suffer mass mortalities due
to climate change (Cebrian et al. 2012). When the invasive alga dominates the benthic
9
communities, it causes ecosystem homogenization, reducing species diversity and structural
complexity (Katsanevakis et al. 2014a; Morri et al. 2019). Eutrophication favours the invasiveness
of C. cylindracea and enhances its competitiveness over native macroalgae (Gennaro and Piazzi,
2014; Gennaro et al. 2015). C. cylindracea dense turf formations trap sediments and may create
anoxic conditions underneath the surface that negatively affect biodiversity (Piazzi et al. 2007;
Klein and Verlaque 2008). This sediment trapping can lead to the replacement of rocky-substrate
fauna by soft-substrate fauna, as observed with Caulerpa taxifolia (M.Vahl) C.Agardh and
Lophocladia lallemandii (Montagne) F.Schmitz (decrease in crustaceans, echinoderms, and
mollusks, and increase in polychaetes). When trapping relevant quantities of sediment, C.
cylindracea can reduce the fitness of seagrass meadows (Ceccherelli and Campo 2002; Dumay et
al. 2002; Najdek et al. 2020). Moreover, turfs of the invasive alga introduce major compositional
changes in the established macrofaunal and macrophytic communities in seagrass meadows
(Argyrou et al. 1999; Klein and Verlaque 2011). In sandy habitats, the invasive alga may also
induce negative effects through competition on its native congeneric Caulerpa prolifera (Forsskål)
J.V. Lamouroux (Balestri et al., 2018). Posidonia oceanica beds resist C. cylindracea invasion due
to the light restrictions caused by dense and healthy meadows (Marín-Guirao et al. 2015;
Bernardeau-Esteller et al. 2015). On the other hand, disturbed seagrass beds and “matte morte”
areas (remnants of former Posidonia oceanica meadows) are vulnerable to the invasion of C.
cylindracea (Katsanevakis et al. 2010; Ceccherelli et al. 2014; Gribben et al. 2018). In contrast, C.
cylindracea can also have positive impacts on seagrasses (Ceccherelli and Campo 2002; Pereda-
Briones et al. 2018). Shoot density of Zostera noltei Hornemann has been observed to increase in
the presence of C. cylindracea while the invasive alga may also facilitate survival and growth of
Posidonia oceanica seedlings, due to the dense turf mats that provide the appropriate habitat for
the angiosperm germlings to flourish (Ceccherelli and Campo 2002; Pereda-Briones et al. 2018).
Rizzo et al. (2020) showed that the presence of C. cylindracea could alter quantity, biochemical
composition, and nutritional quality of organic detritus and change the status of the benthic
ecosystem functioning. In the presence of the invasive chlorophyte in soft bottom ecosystems,
sediment microbial communities exhibit higher metabolic activity (Rizzo et al. 2017b). Organic
matter has been estimated to be higher in soft bottom habitats invaded by C. cylindracea (Pusceddu
et al. 2016; Rizzo et al. 2017b). The observed effects could be context-dependent. At low sediment
deposition rates, C. cylindracea affects the resident meiofaunal communities positively, whereas
at high to medium rates the effect is negative (Rizzo et al. 2020). C. cylindracea is consumed by
various herbivorous and omnivorous species in the Mediterranean (Box et al. 2009a; Cebrian et
al. 2011; Terlizzi et al. 2011; Felline et al. 2017; Noè et al. 2018a, Santamaría et al. 2021a). The
omnivorous fish Diplodus sargus (Linnaeus, 1758) and the herbivorous echinoid Paracentrotus
lividus (Lamarck, 1816) are among the species that have incorporated C. cylindracea in their diet
(Terlizzi et al. 2011; Tomas et al. 2011a). However, C. cylindracea consumption affects the fitness
of the two species altering their physiology and behaviour (Tomas et al. 2011a; Tejada et al. 2013;
Gorbi et al. 2014; Magliozzi et al. 2017, 2019; Vitale et al. 2018; Vega Fernández et al. 2019;
Miccoli et al. 2021). This reaction is probably due to the toxic activity of C. cylindracea secondary
metabolites such as caulerpin, caulerpenyne and caulerpicin. The invasive alga has exhibited high
tolerance against herbivory pressure, due to its regenerative abilities (Bulleri and Malquori 2015).
A strong tolerance has also been observed in highly invaded areas and P. lividus, the major
10
Mediterranean herbivore, exhibited the ability to confine the invasive alga only when it is present
in low densities (Cebrian et al. 2011). However, when herbivory acts synergistically with the biotic
resistance of a healthy Mediterranean macroalgal community the invasion success of C.
cylindracea can be significantly limited (Santamaría et al. 2021b). High caulerpin accumulations
to tissues of the native herbivorous fish Sarpa salpa (Linnaeus, 1758) have been associated with a
potential negative toxic effect on human health after consumption (Turhan and Cavas 2019). The
provision of ecosystem services such as food, biotic materials, lifecycle maintenance, symbolic
and aesthetic values, recreation, and cognitive benefits has been negatively impacted by the
invasion of C. cylindracea (Salomidi et al. 2012; Otero et al. 2013; Katsanevakis et al. 2014a).
Still, C. cylindracea could reduce sediment erosion contributing thus to coastal protection
(Hendriks et al. 2010; Katsanevakis et al. 2014a). Nevertheless, due to the seasonality of its fronds
surface and weak rhizome structure, the level of coastal protection under the most harsh winter
weather conditions may not be effective. Hence, the replacement of seagrass beds with Caulerpa
meadows would probably introduce major changes in an area’s annual sediment dynamics
(Hendriks et al. 2010). Furthermore, C. cylindracea has been assessed as a potentially promising
resource for the development of functional food, agronomic and pharmaceutical products (De
Souza et al. 2009; Cavas and Ponhert 2010; Stabili et al. 2016a; Agirbasli and Cavas 2017; Vitale
et al. 2018; Alburquerque et al. 2019). C. cylindracea extracts have been associated with
antioxidant, antibacterial, antiviral and anticancer activities (e.g. Vairappan 2004; Cavas and
Yurdakoc 2005; Vieira Macedo et al. 2012; Mehra et al. 2019).
Caulerpa taxifolia (M.Vahl) C.Agardh
Caulerpa taxifolia is an invasive chlorophyte first discovered in 1984 off the Oceanographic
Museum of Monaco (Verlaque et al. 2015). The invasive alga spread rapidly and aggressively
throughout the western and central Mediterranean and the Adriatic Sea. C. taxifolia forms seasonal
dense meadows, on various types of rocky or sandy substrata that may outcompete hard substrate
algae and under certain conditions Mediterranean native seagrasses (Santini-Bellan et al. 1996; de
Villèle and Verlaque 1995; Ceccherelli and Cinelli 1997; Dumay et al. 2002; Piazzi and Cinelli
2003; Cattaneo-Vietti 2018). In rocky habitats invaded by C. taxifolia, macroalgal communities
have been reported to undergo a drastic impoverishment (Verlaque and Fritayre 1994). Native
species gradually disappear after the invasion, with the first to undergo a drastic decline to be
fleshy macroalgae alongside with their epiphytes, then filamentous macroalgae and finally
crustose algae. The negative effect maximizes at the seasonal peak of C. taxifolia development at
the end of summer and autumn. The newly formed habitats of C. taxifolia introduce a significant
degradation to the previously established communities of benthic invertebrates and fish (Otero et
al. 2013; Deudero et al. 2014; Katsanevakis et al. 2014a; Montefalcone et al. 2015b; Cheminée et
al. 2016; Cvitković et al. 2017). On the contrary, novel habitats created by C. taxifolia act as
shelters for other organisms (Gianguzza et al. 2013; Katsanevakis et al. 2014a; Cvitković et al.
2017). Eventually, it was documented that the loss of seagrasses due to C. taxifolia competition is
possible in disturbed meadows, but healthy seagrass beds are probably not affected by C. taxifolia
invasion (de Villèle and Verlaque 1995; Jaubert et al. 1999, 2003; Ceccherelli and Sechi 2002;
Glasby 2013). Still, C. taxifolia, negatively affects the provision of ecosystem services derived
from invaded biotopes, such as the provision of biotic materials, coastal protection, cognitive
11
benefits, recreation and tourism, symbolic and aesthetic values, and life cycle maintenance
(Salomidi et al. 2012; Otero et al. 2013; Katsanevakis et al. 2014a). Even more, these dense algal
aggregations entangle in fishing nets negatively affecting the activity of fishers (Bianchi et al.
2019a). On the other hand, C. taxifolia sediment stabilization and trapping contributes to coastal
protection by reducing sediment erosion (Hendriks et al. 2010; Katsanevakis et al. 2014a).
Nevertheless, due to the seasonality of its fronds surface and weak rhizome structure, the level of
coastal protection under the most harsh winter weather conditions may not be effective (Hendriks
et al. 2010). Hence, the replacement of seagrass beds with Caulerpa meadows would probably
introduce major changes in an area’s annual sediment dynamics (Hendriks et al. 2010). Underneath
the matte of the invasive alga, anoxic conditions may occur. Rocky substrate fauna is replaced by
soft substrate fauna, as observed for instance with polychaetes (Santini-Bellan et al. 1996;
Francour et al. 2009; M. Verlaque pers. obs.). C. taxifolia rapidly expanded its range in the
Mediterranean until 2000 but afterwards its dispersal rate slowed and followed a regressive phase
even at areas where it had aggressively invaded (Montefalcone et al. 2015a; Chefaoui and Varela-
Álvarez 2018). C. taxifolia produces substances that have been associated with anticancer, anti-
proliferative, anti-microbial, anti-herpetic and anti-viral properties (e.g. Barbier et al. 2001; Cavas
and Pohnert 2010; Sfecci et al. 2017; Mehra et al. 2019; Bayro et al. 2021).
Caulerpa taxifolia var. distichophylla (Sonder) Verlaque, Huisman & Procaccini
The chlorophyte Caulerpa taxifolia var. distichophyllais a variant of Caulerpa taxifolia that
originates from southwestern Australia, and a recent invader of the Mediterranean Sea that was
first found in Syria in 2003 (Bitar et al. 2017), then in Turkey in 2006 (Cevik et al. 2007; as C.
taxifolia) and in Sicily in 2007 where large drift biomass washed ashore (Cormaci and Furnari
2009; Meinesz et al. 2010; Jongma et al. 2013). The invasive chlorophyte is rapidly expanding its
range in the central and eastern Mediterranean (Musco et al. 2014; Aplikioti et al. 2016; Picciotto
et al. 2016; Ellul et al. 2019). In Sicily the species exhibits the ability to monopolize the bottom at
dead Posidonia oceanica matte (Musco et al. 2014). Furthermore, invaded P. oceanica meadows
revealed an opportunistic polychaete assemblage compared to a community structured by more
sensitive polychaetes to meadows unaffected by the invasion of C. taxifolia var. distichophylla
(Musco et al. 2014). In Rhodes Island C. taxifolia var. distichophylla formed monospecific mats
at a variety of habitats between 9 and18 m depth (Aplikioti et al. 2016). Furthermore, it was
observed to cover Cystoseira sensu lato communities, adding another pressure to the overstressed
and declining communities of perennial algae (Thibaut et al. 2005, 2015; Aplikioti et al. 2016).
C. taxifolia var. distichophylla has exhibited low palatability by the major Mediterranean grazer,
the echinoid Paracentrotus lividus (Noè et al. 2018a). Still, the chlorophyte was consumed by the
echinoid when offered as a mixture with its congeneric Caulerpa cylindracea (Noè et al. 2018a).
Nevertheless, the mixture of the two alien chlorophytes enabled a synergistic anti-grazing effect
on the fitness of P. lividus due to their toxic metabolites that changed the behavior of the echinoid
grazer (Vega Fernández et al. 2019). C. taxifolia var. distichophylla can be a nuisance to
commercial fishers, entangling in their nets (Musco et al. 2014; Verlaque et al. 2015). Maritime
traffic and the continuous warming of the Mediterranean basin will probably facilitate further
spreading of C. taxifolia var. distichophylla (Mannino et al. 2019).
12
Cladophora patentiramea (Montagne) Kützing
Cladophora patentiramea is a filamentous green alga, originating from the Indo-Pacific Ocean. It
has been introduced in the Mediterranean Sea through the Suez Canal and reported from Cyprus
and Lebanon (Verlaque 1994; Bitar et al. 2017). The species can form massive blooms in shallow
waters, forming “nets” that can hinder bathing (Verlaque et al. 2015). Its impact on biodiversity
has not been studied yet.
Codium arabicum Kützing
The Indo-Pacific Codium arabicum was first observed in 2007 in the Mediterranean Sea from
Haifa Bay, Israel (Hoffman et al. 2011; Verlaque et al. 2015). C. arabicum was the dominant algal
species in seaweed drifts at Haifa bay shores, during the summer of 2009 (Hoffman et al. 2011).
These drifts negatively affect ecosystem services such as aesthetic values, recreation and tourism.
The appearance and proliferation of the invasive seaweeds C. arabicum, C. parvulum, Galaxaura
rugosa, and Stypopodium schimperi alongside with marine pollution, have been associated with
the replacement of native macroalgal species at Haifa Bay (Hoffman et al. 2008, 2011).C.
arabicum extracts are known to have antibacterial, anticancer and antifungal activities (e.g. Sheu
et al. 1995; Padmakumar and Ayyakkannu 1997; El Zawawy et al. 2020).
Codium fragile subsp. fragile (Suringar) Hariot
The green alga Codium fragile subsp. fragile, previously referred to as Codium fragile subsp.
tomentosoides, is a global invader that originates from NW Pacific. In the Mediterranean it was
first reported in 1946 in southern France (Verlaque 1994; Verlaque et al. 2015) and observations
indicate that its abundance peaked about a decade after its first discovery, and then its numbers
followed a regressive phase (Schaffelke and Hewitt 2007). Nonetheless, the species is still
widespread in the basin and has established populations that often thrive in subtidal rocky bottoms,
forming dense sponge-like fronds of low height that become a major structural element of the
invaded habitat and dominate the macroalgal community (Tsiamis 2012; Otero et al. 2013; Tsiamis
et al. 2013; Katsanevakis et al. 2014a). As a structural engineer C. fragile favours a variety of
organisms that settle among its bushy structures, such as mussels (Bulleri et al. 2006; Katsanevakis
et al. 2014a). Nevertheless, its dense erect formations increase sedimentation and inhibit light from
reaching the substrate, thus shading and negatively affecting various other benthic organisms
(Katsanevakis et al. 2014a). C. fragile causes negative impacts on food provisioning ecosystem
services, due to fouling on aquaculture structures and even on fishing nets (Streftaris and Zenetos
2006; Katsanevakis et al. 2014a; Orlando-Bonaca et al. 2021). Additional ecosystem services
negatively impacted by the invasive green alga (Katsanevakis et al., 2014a) are biotic materials,
cognitive benefits, recreation, symbolic and aesthetic values, and life cycle maintenance (Salomidi
et al. 2012). Also, when abundant it can produce massive drifts that seasonally wash up on beaches,
having a negative impact on aesthetic values and recreation and tourism (Streftaris and Zenetos
2006; Katsanevakis et al. 2014a). As dimethylsulfoniopropionate (DMSP) producer it can have a
positive impact on climate regulation but also a negative impact on air quality regulation
(Katsanevakis et al. 2014a). C. fragile has been assessed as a valid biomonitoring species for high
trace element concentrations in the marine environment (Malea et al. 2015). Extracts of C. fragile
13
have cytotoxic, antioxidant and antimicrobial activities (e.g. Kim et al. 2006; Kimiya et al. 2008;
Koz et al. 2009; Kim et al. 2013).
Codium parvulum (Bory ex Audouin) P.C.Silva
The green seaweed Codium parvulum was first reported from the eastern Mediterranean Sea in
2004, on the northern shores of Israel (Israel et al. 2010; Verlaque et al. 2015). Massive amounts
of C. parvulum have repeatedly been washed ashore on the sandy shores of Israel during winter
and spring, degrading the aesthetic value of the coast but also disturbing human activities such as
fishing and touristim (Israel et al. 2010; Hoffman et al. 2011; Verlaque et al. 2015). C. arabicum,
C. parvulum, Galaxaura rugosa, and Stypopodium schimperi have become abundant after their
appearance in Haifa Bay and alongside with marine pollution, have been associated with a negative
effect on native macroalgal species (Hoffman et al. 2008, 2011). The invaded geographical range
of C. parvulum is restricted to the Eastern Mediterranean, where it can form abundant populations
at specific localities (Bitar et al. 2017).
Halimeda incrassata (J.Ellis) J.V.Lamouroux
The tropical chlorophyte Halimeda incrassata is a recent invader of the western Mediterranean,
first recorded in 2011 off Mallorca Island (Alós et al. 2016). By 2015, this invasive chlorophyte
had rapidly spread and occupied extensive surfaces of a monitored marine protected area at the
Mallorca Island (Alós et al. 2016). While it is not a significant competitor of Posidonia oceanica,
H. incrassata has been indicated to compete with native macroalgae such as Dasycladus
vermicularis (Scopoli) Krasser and activate oxidative stress responses (Sureda et al. 2017). As an
ecosystem engineer and a creator of novel habitats, its impacts on native fauna are both negative
and positive (Alós et al. 2018; Vivó-Pons et al. 2020). Novoa et al. (2011) and Costa-Mugica et
al. (2012) demonstrated the antioxidant capacities of the lyophilized aqueous extract of H.
incrassata, which are related with the phenolic components of the extract.
Rhodophyta
Acrothamnion preissii (Sonder) E.M.Wollaston
Acrothamnion preissii is a tropical rhodophyte of Indo-Pacific origin that was first reported in the
Mediterranean Sea in 1969 from Italy (Livorno). It has become invasive in many localities,
particularly in the western part of the basin (Verlaque et al. 2015). A. preissii forms dense turf
communities that overgrow native sessile species and deteriorate the state of ecologically
important habitats such as coralligenous assemblages, maërl beds, and Posidonia oceanica
meadows (Ferrer et al. 1994; Piazzi et al. 1996, 2001, 2015; Piazzi and Cinelli 2000, 2001, 2003;
Linares et al. 2012). The invasive rhodophyte acts as a habitat/ecosystem engineer by negatively
affecting the substrate available for other benthic organisms to settle on, and reducing light
availability for other photosynthetic species (Katsanevakis et al. 2014a; Smith et al. 2014). It can
also grow on P. oceanica leaves with a negative shading effect on this keystone species
(Katsanevakis et al. 2014a). Dense turfs of A. preissii have a high surface to volume ratio that
14
reflects in high nutrient uptake rates, resulting in a negative competitive impact for resources
(nutrients) on coarser algae (Katsanevakis et al. 2014a). A. preissi outcompetes native macroalgae
for space and resources; Accordingly, ecosystem services provided by sublittoral macroalgal
communities such as food, biotic materials, cognitive benefits, recreation and tourism, symbolic
and aesthetic values, and life cycle maintenance (Salomidi et al. 2012) are negatively impacted
(Katsanevakis et al. 2014a). This invasive rhodophyte can affect fisheries by clogging on fishing
nets and thus reducing available catch (Cinelli et al. 1984). A. preissii could also have a negative
effect on microorganisms because it is never epiphytized, probably due to the production of
antifouling and repellent substances (M. Verlaque, unpublished observation). One of the major
Mediterranean herbivores, the sea urchin Paracentrotus lividus (Lamarck, 1816), avoids the
consumption of A. preissii (Tomas et al. 2011a). It is common for rhodophytes to produce repellent
secondary metabolites as is the case for A. preissi and the typical highly refractive cells it produces
(gland cells) that either release extracellular compounds or serve as storage cells (Guiry and Guiry
2021).
Agarophyton vermiculophyllum (Ohmi) Gurgel, J.N.Norris & Fredericq
The perennial red alga Agarophyton vermiculophyllum originates from the Sea of Japan and was
first reported from the Mediterranean in the Po Delta lagoons (N Adriatic) in 2008 (Sfriso et al.
2010; Verlaque et al. 2015, as Gracilaria vermiculophylla (Ohmi) Papenfuss). A recent molecular
study distinguished new lineages within the Gracilariales order and established Agarophyton as a
new genus within the Gracilariaceae (Gurgel et al. 2018). A. vermiculophyllum is abundant in the
Po Delta, Venice lagoon, and the Emilia-Romagna region, where it can replace native macroalgae
(Sfriso et al. 2012; Sfriso and Marchini 2014). Still, the replacement of Ulvaceae has been
beneficial, since it has prevented the anoxic crisis caused by massive Ulva spp. mortalities that
occured when water temperatures exceeded 25–26°C (Sfriso and Marchini 2014; Sfriso et al. 2018,
2020). It is an important structural ecosystem engineer that introduces habitat complexity, as it has
been found to support a diverse macrofaunal community and to act as shelter and nursery area
(Katsanevakis et al. 2014a; Munari et al. 2015; Haram et al. 2020). In the USA, A.
vermiculophyllum dense canopies can change the hydrodynamic state of tidal flats introducing
major changes on biodiversity and ecosystem functioning (Volaric et al. 2019). In high densities
it can create thick mats on soft bottoms that trap sediment and improve water transparency, but
eventually it can prevent light from reaching understory communities and reduce nutrient
availability for other macroalgae (Katsanevakis et al. 2014a). The species has potential for
commercial exploitation for uses in agar, sugar and alcohol production (Sousa et al. 2012; Sfriso
and Marchini 2014; Sfriso et al. 2016). It has been successfully farmed in the Venice Lagoon in a
pilot study that tested its biomass production for commercial cultivation (Sfriso and Sfriso 2017).
It has the ability to act as a biofilter and reduce nitrogen load from the water in fish farming (Shin
et al. 2020). Furthermore, A. vermiculophyllum fouls fishing gear, cultivated shellfish and
aquaculture structures, causing a negative impact on food provision (Katsanevakis et al. 2014a).
Seasonally abundant drifting algae have been observed to wash up on nearby coasts, a negative
impact on aesthetic values and on recreation (Katsanevakis et al. 2014a).
Antithamnion nipponicum Yamada & Inagaki
15
The NW Pacific Antithamnion nipponicum was firstly recorded in the Mediterranean in 1988 in
the Thau Lagoon (France) (Cho et al. 2005; Verlaque et al. 2015). The invasive rhodophyte is often
misidentified as Antithamnion pectinatum (Montagne) Brauner. Antithamnion nipponicum is also
introduced along the inner shores of the city of Venice, settling extensively on vertical substrates
from +0.2 to −8 m (Curiel et al. 1998, as A. pectinatum). The species occupies various types of
substrate at a wide depth range, becomes an epiphyte on other algae and shellfish, and competes
with other filamentous algae such as Antithamnion cruciatum (C. Agardh) Nägeli, Dasya spp.,
Aglaothamnion sp. and Spermothamnion sp. (Curiel et al. 1998). Antithamnion nipponicum is
widespread and can be found both in turbid and clear waters (Curiel et al. 1998). Furthermore, it
is considered a pest for shellfish aquaculture in coastal lagoons as it fouls on their facilities
(Streftaris and Zenetos, 2006). It achieves effective clonal reproduction by fragmentation and has
developed effective anti-grazing mechanisms (Rueness et al 2007).
Asparagopsis armata Harvey
The alien Rhodophyte Asparagopsis armata was probably introduced in the western
Mediterranean Sea during the 1920s and is now widely distributed throughout the basin (Verlaque
et al. 2015; Zanolla and Andreakis 2016). A. armata has a heteromorphic life-history with a small
filamentous tetrasporophytic phase and a mid-size erect gametophytic phase. Gametophytes create
an erect algal cover that dominates rocky bottoms and negatively affects benthic communities and
native species (Smith et al. 2014; Longobardi et al. 2017; Rueda et al. 2021; Orlando-Bonaca et
al. 2021). In western Portugal, rock pools that were invaded by the invasive rhodophyte exhibited
a less complex community with less macroalgal species than non-invaded rock pools (Silva et al.
2021a). The articulated coralline Ellisolandia elongata (J. Ellis & Solander) K.R. Hind & G.W.
Saunders was the most impacted macroalgal species suffering significant biomass losses (Silva et
al. 2021a). A. armata can also be an epiphyte on native macroalgae during seasonal blooms
(Orlando-Bonaca et al. 2017). When overgrowth is achieved by dense A. armata assemblages, less
light and nutrients reach the host algae (Katsanevakis et al. 2014a). Furthermore, the
tetrasporophyte, by its high ratio of surface to volume, has a greater potential for rapid uptake of
nutrients in comparison to their host algae, reducing nutrients availability for them (Katsanevakis
et al. 2014a). At the Atlantic coasts of the Iberian Peninsula, a comparison between E. elongata
and A. armata isopod communities revealed higher abundance and species richness in stands of
the native E. elongata (Guerra-García et al. 2012). On the other hand, specific taxonomic groups
such as epiphytic microalgae, crustaceans and polychaete consumers, are being favoured by the
habitats created by the invasive rhodophyte (Pacios et al. 2011; Soler-Hurtado and Guerra-García
2011; Guerra-García and Sánchez-Moyano 2013; Ricevuto et al. 2015; Moncer et al. 2017). A.
armata releases toxic compounds in the invaded areas where it establishes, in order to prevail. The
fitness of the common prawn Palaemon elegans Rathke, 1836 and the marine snail Steromphala
umbilicalis (da Costa, 1778) are negatively affected when they interact with A. armata due to toxic
exudates produced by the invasive rhodophyte (Silva et al. 2021b). On the other hand, A. armata
has positive effects on water purification, and its tetrasporophytic phase is an efficient biofilter in
fish aquaculture (Katsanevakis et al. 2014a). However, it can be a nuisance to commercial fishers
since it is clogging their nets (Verlaque et al. 2015). It can also overgrow oysters (Mineur et al.
2007). Various algal extracts have been derived from A. armata for medical research and for
16
breeding use (fish farming) due to anti-bacterial, anti-cancer, anti-protozoan and antioxidant
properties (Bansemir et al. 2006; Paul et al. 2006; Genovese et al. 2009; Zubia et al. 2009; Zbakh
et al. 2012; Bouhlal et al. 2013; Rhimou et al. 2013; Castanho et al. 2017). A. armata is generally
avoided by herbivores, due to its anti-grazing mechanisms in contrast to various Mediterranean
native erect fleshy macroalgae (Sala and Boudouresque 1997), but still at some occasions native
herbivores have been recorded to include it on their diet (Vergés et al. 2008; Martínez-Crego et al.
2020). In areas where the species dominates macroalgal communities, large quantities of dead
plants (both phases) may accumulate on the bottom and on beaches, with a negative effect on
aesthetic values and recreation and tourism (Katsanevakis et al. 2014a; M. Verlaque pers. obs.).
Asparagopsis taxiformis (Delile) Trevisan de Saint-Léon
The rhodophyte Asparagopsis taxiformis was first described in 1813 from the eastern
Mediterranean in Alexandria, Egypt (Katsanevakis et al. 2014a). Initially A. taxiformis had
colonized only a part of the eastern Mediterranean Sea, but during the 1990s the species’ records
throughout the basin multiplied and various established invasive populations were reported all over
the Mediterranean (Tsiamis et al. 2013). This is probably because A. taxiformis is composed of
multiple genetically distinct lineages of uncertain taxonomic status (Andreakis et al. 2007, 2009,
2016; Zanolla et al. 2014, 2015) that have invaded the Mediterranean at different phases.
Molecular analyses revealed that two strains exist in the Mediterranean Sea, a strain of Indo-Pacific
origin, which is quite widespread in the western basin, but also in the Adriatic Sea and Greece,
and a less invasive strain of tropical Atlantic origin, found in the eastern Mediterranean Sea
(Andreakis et al. 2007, 2009; Verlaque et al. 2015). A. taxiformis has a heteromorphic life cycle
with a small, tufted tetrasporophyte (often referred to as the Falkenbergia hillenbrandii (Bornet)
Falkenberg phase) morphologically quite difficult to separate from the tetrasporophyte of A.
armata (Zanolla et al. 2014). The tetrasporophytes have a high ratio of surface to volume which
gives them greater potential for rapid uptake of nutrients in comparison to their host algae, a
negative impact on ocean nourishment (Katsanevakis et al. 2014a). Gametophytes of A. taxiformis
form dense seasonal upright formations that regularly monopolize rocky reef macroalgae
communities and are negatively correlated with native algae (Altamirano et al. 2008; Tsiamis
2012; Lodola 2013; Tsiamis et al. 2013; Zanolla et al. 2018). Mancusoa et al. (2021) compared
biomass of the native fucoid Ericaria brachycarpa (J. Agardh) Molinari & Guiry and the invasive
rhodophyte A. taxiformis and also species composition and structure of the epifaunal assemblages
that inhabit the provided habitat by the two macroalgae. A. taxiformis biomass proved to be 90%
lower than that of the native canopy forming E. brachycarpa. Abundance, species richness, and
Shannon-Wiener diversity index of the epifaunal assemblages decreased significantly in the
invasive red alga formations in comparison to E. brachycarpa habitat. Furthermore, Navarro-
Barranco et al. (2018) showed that A. taxiformis, acting as a structural ecosystem engineer, hosted
an impoverished epifaunal assemblage in comparison to the native Halopteris scoparia (Linnaeus)
Sauvageau (Sphacelariales, Ochrophyta). The latter study also revealed a biotic homogenization
of the epifaunal assemblages associated with A. taxiformis. A. taxiformis thrives under high
substrate complexity, coexisting with the invasive chlorophyte Caulerpa cylindracea Sonder,
whereas on homogenous platforms the invasive rhodophyte is absent even in experiments where
other algae were removed, depending on heterogeneity in environmental conditions (Tamburello
et al. 2013). Negative interaction effects were also found between A. taxiformis and the coral
Astroides calycularis (Pallas, 1766) with an observed significant increase of the rhodophyte’s
bioactivity against the marine bacterium Aliivibrio fischeri (Beijerinck, 1889) that was interpreted
17
as a defensive response or as allelopathic offense (Greff et al. 2017). A. taxiformis has a potential
to be farmed to produce economically valuable secondary metabolites (bromoform and
dibromoacetic acid) (Mata et al. 2011, 2012). When consumed by livestock animals the bioactive
compound bromoform can reduce enteric CH4 production (Abbott et al. 2020). A. taxiformis crude
ethanolic extracts have shown a promising activity for new drug treatments against Leishmania
species, a genus of protozoan parasites that cause the disease Leishmaniosis to animals and humans
(Genovese et al. 2009; Vitale et al. 2015). Furthermore, its antioxidant ability further establishes
A. taxiformis as a prominent and novel resource for pharmaceutical compounds (Mellouk et al.
2017). Asparagopsis taxiformis ethanol extracts have exhibited antibacterial activity against
pathogenic fish and shellfish bacteria and have been assessed suitable for use in aquaculture
(Genovese et al. 2012; Marino et al. 2016). Furthermore, aqueous extracts of A. taxiformis contain
high concentrations of cytokinins and have been assessed as prominent for agricultural use,
improving in vitro plant regeneration and micropropagation of apricot hypocotyl slices
(Alburquerque et al. 2019).
Bonnemaisonia hamifera Hariot
The NW Pacific Bonnemaisonia hamifera has been introduced in the European seas for more than
a century, but much later in the Mediterranean Sea (first observation of sporophyte in 1909 in
Tunisia: Petersen 1918; Verlaque et al. 2015). B. hamifera has a heteromorphic life history with
alternation of erect gametophytes and filamentous tetrasporophyte. The success of B. hamifera is
most likely due to its chemical defenses, through the production of potent brominated compounds
(Enge et al. 2012), deterring herbivores and reducing the growth and survival of seaweed
propagules and microbes that settle on its surface (Nylund et al. 2005; Enge et al. 2012, 2013;
Svensson et al. 2013). In other European seas the species has been proven to determine negative
impacts as an ecosystem engineer by forming dense epiphytic growth on host algae (e.g. Johansson
et al. 1998), and by preventing competing algae to colonize (Svensson et al. 2013). Its high surface
to volume ratio and associated greater potential for rapid uptake of nutrients in comparison to their
host algae is perceived as a negative impact on ocean nourishment (Katsanevakis et al. 2014a).
Although B. hamifera, mainly the filamentous tetrasporophyte, is well established throughout the
Mediterranean Sea, no pullulation has been reported and there is no documentation of caused
impacts on Mediterranean biodiversity or ecosystem services. The excessively high temperatures
and salinities of the Mediterranean Sea could be unfavourable to the development of this cold
temperate species.
Galaxaura rugosa (J. Ellis & Solander) J.V. Lamouroux
The invasive rhodophyte Galaxaura rugosa was first reported in 1990 in Syria (Mayhoub, 1990)
as Galaxaura lapidescens (J. Ellis & Solander) J. V. Lamouroux and has become a common
macroalgal species along the coasts of the eastern Levantine Sea (Hoffman et al. 2007; Hoffman
and Dubinsky 2010; Verlaque et al. 2015). It flourishes on rocky habitats during the summer, and
it has probably competitively displaced native macroalgae (Hoffman and Dubinsky 2010),
possibly indirectly assisted through avoidance by the dominant herbivores (siganid fishes) (Peleg
et al. 2019), as it was similarly observed in Australia (Mantyka and Bellwood 2007). The
appearance and proliferation of the invasive seaweeds Codium arabicum, C. parvulum, G. rugosa,
and Stypopodium schimperi in Haifa Bay have been associated with a negative effect on native
18
macroalgal species alongside with marine pollution (Hoffman et al. 2008, 2011). During the years
2004–2008, massive algal blooms led to large quantities of algal drift washing ashore the coasts
of northern Israel (Israel et al. 2010). These massive biomass quantities of algal material, mostly
composed of fragmented G. rugosa (30% of the total amount, Hoffman and Dubinsky 2010), C.
parvulum, and S. schimperi, discouraged winter sport activities and degraded the aesthetic value
of the coast (Hoffman et al. 2011). On the other hand G. rugosa extracts have exhibited
antimicrobial, antioxidant, antifungal and anticancer activities (Al-Enazi et al. 2018).
Ganonema farinosum (J.V. Lamouroux) K.-C. Fan & Y.-C. Wang
Widespread in the Mediterranean Sea, Ganonema farinosum is yet of unknown origin and
classified as cryptogenic by Zenetos et al. (2010) and Veraque et al. (2015). Currently there are no
documented impacts on biodiversity or ecosystem services by G. farinosum within the
Mediterranean basin (Verlaque et al. 2015). However, high densities of the species have been
recorded in the southern Aegean Sea, becoming occasionally the dominant alga in overgrazed
shallow rocky reefs, probably because grazers avoid consuming it (Thessalou-Legaki et al. 2012;
Tsiamis 2012; K. Tsirintanis, pers. obs.). G. farinosum has been found to be positively affected by
overgrazing by siganids as it seems to be avoided by these herbivores and at the same time other
macroalgae are rapidly depleted by the invasive herbivores (Gerovasileiou et al. 2017; Dimitriadis
et al. 2021).
Grateloupia turuturu Yamada
The Japanese red alga Grateloupia turuturu was first detected in the Mediterranean Sea in the
Thau Lagoon in 1982 (Riouall et al. 1985, as G. doryphora; Verlaque et al. 2015) and has spread
all over the Mediterranean basin. This alien rhodophyte was often misidentified as Grateloupia
doryphora (Montagne) M.Howe (Gavio and Fredericq 2002; Zenetos et al. 2011; Cecere et al.
2011). G. turuturu is a large perennial rhodophyte with blades up to 80 cm long and 18 cm wide,
and it acts as a structural engineer that favours a variety of organisms, while reducing the presence
of others (Cecere et al. 2011; Katsanevakis et al. 2014a; Smith et al. 2014). G. turuturu is an alien
rhodophyte with the potential to cause significant environmental impacts. These impacts range
from alterations in ecosystem functioning to modifications of the biological communities
(Mathieson 2010 and references therein). But most evidence of impact is based on low inferential
strength of evidence with a luck of experimental studies. As only the encrusting base is perennial,
the role of engineer species only operates during the seasonal development of erect fronds. Fronds
of G. turuturu are rarely epiphytized and do not show grazing marks (Jones and Thornber 2010;
M. Verlaque, pers. obs.). Within its invaded range in the Mediterranean, it has a negative impact
on food provision services by fouling shellfish and aquaculture facilities in coastal lagoons, as well
as by growing on fishing nets (Streftaris and Zenetos 2006; Cecere et al. 2011). G. turuturu has
the potential to be farmed as it is edible, rich in proteins and dietary fibres, and has been recognized
as a prominent nutritional food source (Denis et al. 2010; Kendel et al. 2013). Extracts of G.
turuturu have antibacterial, anti-microfouling, antiprotozoal, antiviral, and UV-protection
activities (Plouguerné et al. 2008; Liu and Pang 2010; García-Bueno et al. 2014; McReynolds
2017).
19
Lophocladia lallemandii (Montagne) F.Schmitz
The red alga Lophocladia lallemandii is an old invader of the Mediterranean Sea, recorded first in
1908 in Greece and Libya (Petersen 1918). The species has colonized the entire basin causing
significant impacts especially on western Mediterranean ecosystems (Zenetos et al. 2010;
Verlaque et al. 2015; Cebrian et al. 2018). L. lallemandii exhibits rapid invasive dispersal, with
the ability to grow on a variety of different habitats and become a dominant species that
homogenizes the substrate by creating dense turf mats (Patzner 1998; Boudouresque and Verlaque
2002; Ballesteros 2006; Cebrian and Ballesteros 2010; Bedini et al. 2011; Katsanevakis et al.
2014a). L. lallemandii invasion constitutes a major threat for the Mediterranean endemic
Posidonia oceanica meadows. The invasive alga invades seagrass meadows as an epiphyte of
seagrass leaves, and forms dense turf mats that reduce the fitness of the meadow, eventually
increasing shoot mortality rates (Ballesteros et al. 2007; Sureda et al. 2008; Marbà et al. 2014; El
Zrelli et al. 2021). The dense mats trap sediment causing reduction of water exchange and anoxic
conditions within the meadow (Ballesteros et al. 2007). L. lallemandi epiphytism on P. oceanica
meadows degrades the invaded habitat and negatively affects invertebrate communities (Deudero
et al. 2010). L. lallemandii turf mats constitute novel habitats that favour a number of species,
while reducing the presence of others (Bedini et al. 2015; Katsanevakis et al. 2014a; Tiberti et al.
2021). Within P. oceanica meadows, the invasive rhodophyte exhibits a high level of epizoism on
the fan mussel Pinna nobilis Linnaeus, 1758 shells with various harmful effects on its
physiological responses (Box et al. 2009b; Cabanellas-Reboredo et al. 2010; Vázquez-Luis et al.
2014; Kersting and García-March 2017). By trapping sediments, L. lallemandii turfs might
contribute positively to coastal protection. However, it constitutes a degradation to this ecosystem
service as the seagrass meadows of P. oceanica are more efficient in this process (Katsanevakis et
al. 2014a). L. lallemandii turfs are structured by thin filaments, which means that they have a high
surface to volume ratio, resulting in high uptake rates of nutrients and carbon dioxide, negatively
affecting coarser macroalgal competitors (Katsanevakis et al. 2014a). The provided ecosystem
services that are impacted at invaded habitats by L. lallemandii turfs are food provision, biotic
materials, coastal protection, cognitive benefits, recreation and tourism, symbolic and aesthetic
values, and life cycle maintenance (Katsanevakis et al. 2014a). L. lallemandii overgrows also
macroalgae and can form its turf mats on their understory communities and habitats (Ballesteros
et al. 2007; Sciberras and Schembri 2007; Bedini et al. 2011; Kersting et al. 2014). Furthermore,
L. lallemandii epiphytizes on the invasive chlorophyte Caulerpa taxifolia which responds by
increasing the caulerpenyne and H2O2 production and the antioxidant enzymes activities as a
defense mechanism against the invasive rhodophyte (Box et al. 2008). L. lallemandi can overgrow
other algae that occur in the interstices of the colonies of the scleractinian Cladocora caespitosa
and possibly introduce negative effects to the understory communities by reducing light
penetration (Kersting et al. 2014). Nevertheless, no lethal effects of the invasive alga were detected
in C. caespitosa colonies (Kersting et al. 2014). A multimarker metabarcoding study showed
important structural changes in community structure in the presence of L. lallemandii, and a
reduction of species richness in the size-fraction below 1 mm (Wangensteen et al. 2018).
Occasionally, blooms of the alga have been washed ashore, causing a decrease of aesthetic value
(Occhipinti-Ambrogi 2002). L. lallemandii produces alkaloids with cytotoxic effects known as
lophocladines, which are responsible for increased antioxidant stress responses when consumed
20
by the Mediterranean primary grazers, such as the herbivorous fish Sarpa salpa and the echinoid
Paracentrotus lividus (Tomas et al. 2011b; Tejada et al. 2013; Quetglas-Llabrés et al. 2020). In
invaded habitats, S. salpa avoids feeding on L. lallemandii and P. lividus is less abundant (Tomas
et al. 2011b; Tejada et al. 2013; Quetglas-Llabrés et al. 2020). In areas of high L. lallemandii
density, P. lividus can limit its seasonal development probably by indirect grazing due to the
consumption of native epiphytes of the invasive rhodophyte (Cebrian et al. 2011).
Polysiphonia morrowii Harvey
The NW Pacific Rhodophyta Polysiphonia morrowii was first reported in the Mediterranean Sea
in Thau lagoon in the western Mediterranean in 1997 (Verlaque 2001; Verlaque et al. 2015), then
it spread and became established in various Mediterranean locations (Zenetos et al. 2010). In the
lagoon of Venice, P. morrowii is very abundant particularly in late winter and spring. It colonizes
the hard substrates of islands, breakwaters, and the city docks along a linear surface of ca. 100 km2
with a mean standing crop of 590 g fw m-2. Overall, the total standing crop was estimated at ca.
517 tonnes fw (Sfriso et al. 2020). Currently there are no documented impacts on biodiversity or
ecosystem services by P. morrowii within the Mediterranean basin. The excessively high
temperatures and salinities of the Mediterranean Sea could be unfavourable to the development of
this cold temperate species.
Womersleyella setacea (Hollenberg) R.E.Norris
The filamentous Indo-Pacific rhodophyte, Womersleyella setacea, was first reported in the
Mediterranean Sea in 1986, and then rapidly spread throughout the basin (Zenetos et al. 2010;
Verlaque et al. 2015). It forms dense turf carpets that monopolize rocky substrates, coralligenous
assemblages, Posidonia oceanica meadows, dead P. oceanica mattes and rhodolith assemblages
(free living calcified Rhodophyta, e.g. maërl beds) on a wide bathymetric range. It can alter benthic
community composition and outcompete native species (Airoldi et al. 1994; Athanasiadis 1997;
Airoldi 2000; Piazzi and Cinelli 2000, 2001, 2003; Streftaris and Zenetos 2006; Sciberras and
Schembri 2007; Battelli and Rindi 2008; Piazzi and Balata 2009;Nikolić et al. 2010; Cebrian and
Rodríguez-Prieto 2012; Petrocelli et al. 2019; Piazzi and Ceccherelli 2020). Dense turf carpets
reduce light penetration and negatively affect the understory communities. In Portugal, invaded
rocky communities by the invasive rhodophyte exhibited lower diversity values in comparison to
non-invaded habitats, and erect macroalgal species were the most affected by the invasion (Piazzi
and Balata 2009). Moreover, dense W. setacea turfs overgrow sponge assemblages and accumulate
sediment in their thin filaments, reducing sponge filter-feeding and eventually impeding sponge
reproduction (de Caralt and Cebrian 2013). Moreover, these thick exotic turf formations constitute
new habitats in the ecosystems where they appear, and studies have shown that they lead to a
significant reduction in the abundance and species richness of mobile macrofaunal assemblages
(Bedini et al. 2015; Katsanevakis et al. 2014a). Coralligenous assemblages are also threatened by
W. setacea since the invasive red alga has a wide bathymetric distribution, from shallow waters to
40 m depth. Overgrowth and increased sedimentation negatively affect the survival rate of
calcified encrusting Rhodophyta and gorgonians, and various other fitness indicators (Ribera and
Boudouresque 1995; Cebrian et al. 2012; Linares et al. 2012; Piazzi et al. 2012; Kružić 2014;
García et al. 2015). Nevertheless, some animals can be favoured by these dense alien turfs, and
21
increase their abundance using them for shelter.Invaded assemblageswith accumulated sediments
showed a lower species richness and a decline of decapods and mollusks populations that was
combined with an increase of polychaetes (Bedini et al. 2015). W. setacea dominance is maintained
throughout seasons and depths, a condition that leads to a loss of temporal diversity in zoobenthic
communities (Antoniadou and Chintiroglou 2007). Its establishment in invaded ecosystems
reduces habitat complexity, introduces homogenization, and the new impoverished state is difficult
to reverse (Gatti et al. 2015). The dense alien turf mats with accumulated sediments also degrade
ecosystems and produce a negative impact on the aesthetic values and on recreation and tourism
(Katsanevakis et al. 2014a). The invasive red alga also entangles in fishing nets, impeding fishing
activities and thus having a negative impact on food provision (Verlaque 1989; Katsanevakis et al.
2014a). The dense turf of the thin filamentous W. setacea has a high surface to volume ratio. Due
to this high ratio, W. setacea has a greater potential for rapid uptake of nutrients and carbon dioxide
in comparison to other seaweeds, a negative impact on the availability of nutrients for coarser algae
and a negative impact on ocean nourishment (Katsanevakis et al. 2014a). Furthermore, the
synergistic action of W. setacea overgrowth and temperature increase have been experimentally
tested to cause extensive necrosis to the hexacoral reef builder Cladocora caespitosa colonies
(Kersting et al. 2015). More ecosystem services are impacted by W. setacea invasion due to
overgrowth and increased sedimentation (Katsanevakis et al., 2014a), and these are food provision,
biotic materials, coastal protection, cognitive benefits, and life cycle maintenance (Salomidi et al.
2012). In the Ligurian Sea (NW Mediterranean Sea) an increase in the sea surface temperature
detonated a shift in the sessile epibenthic communities (Bianchi et al. 2019b). W. setacea became
a dominant species of the reef, probably facilitated by the temperature increase, while at the same
time community complexity was reduced (Bianchi et al. 2019b). Substrate coverage increase of
W. setacea has also been associated with nutrient increase (Bulleri and Piazzi, 2015). Although
repellent secondary metabolites have not been investigated, W. setacea is avoided by most grazers
and is not (or at low levels) epiphytized (Tomas et al. 2011a, b; Cebrian and Rodríguez-Prieto
2012, and references therein). W. setacea has not yet been investigated for any potential
bioactivities (McReynolds 2017).
Tracheophyta
Halophila stipulacea (Forsskål) Ascherson
The Red Sea seagrass Halophila stipulacea was first recorded in 1894 in the Mediterranean Sea
(Rhodes Island) and managed to extend from the Eastern Mediterranean Sea to Tunisia, the
Tyrrhenian Sea (Salerno) and the south Adriatic (Albania) (Gambi et al. 2009; Sghaier et al. 2011;
Verlaque et al. 2015). H. stipulacea creates seagrass meadows that constitute a novel habitat for
the invaded ecosystems (Di Martino et al. 2007; Willette and Ambrose 2012; Katsanevakis et al.
2014a; Smith et al. 2014). In a Tunisian marina, H. stipulacea overgrew a Cymodocea nodosa
(Ucria) Asch. meadow, almost completely displacing the native seagrass within 3 years (Sghaier
et al. 2014). Similarly, in other invaded regions such as the Caribbean Sea, H. stipulacea has
caused negative competitive impacts on native seagrass species (Willete and Ambrosi 2012;
22
Smulders et al. 2017). Nevertheless, like all seagrass meadows, H. stipulacea meadows contribute
to a variety of beneficial ecosystem services such as food provision, water purification and ocean
nourishment as they oxygenate waters and sediments and contribute to nutrient cycling, and coastal
protection by their networks of rhizomes that stabilize sublittoral sediments and prevent erosion
(Salomidi et al. 2012; Katsanevakis et al. 2014a). H. stipulacea meadows should also be perceived
as important carbon sinks, higher than C. nodosa meadows and comparable even to Posidonia
oceanica beds, even though more susceptible to remineralization (Apostolaki et al. 2019;
Wesselmann et al. 2021). A comperative study in Sicily, focused on species composition between
algal assemblages of a mixed H. stipulacea and Caulerpa cylindracea meadow and two contiguous
P. oceanica and C. nodosa meadows, showed significant differences in species composition, with
several species exclusive of the H. stipulacea meadow (Di Martino et al. 2006). Nevertheless,
Mediterranean seagrasses host more diverse assemblages of epiphytic flora (Rindi et al. 1999).
The herbivorous fish Sarpa salpa and the green turtle Chelonia mydas (Linnaeus, 1758) have
incorporated the alien seagrass in their diet (Di Genio et al. 2021; Palmer et al. 2021). The projected
tropicalization of the Mediterranean Sea and the seawater temperature increase is expected to
signal a further spread of the alien seagrass within the Mediterranean (Nguyen et al. 2020). H.
stipulacea has been assessed as an effective biomonitoring species of trace elements such as As,
Cd, Cu, Mn, Ni and Zn in sediments (Bonanno and Raccuia 2018). Furthermore, the alien seagrass
contains substances that have shown antioxidant, hypolipideamic and cytotoxic activities, with a
potential for exploitation for pharmaceutical products (Kandemir-Cavas et al. 2019; Bel Mabrouk
et al. 2020; Sansone et al. 2020).
Porifera
Paraleucilla magna Klautau, Monteiro & Borojevic, 2004
Paraleucilla magna is the only alien calcareous sponge introduced in the Mediterranean Sea,
(Longo et al. 2007). Despite the fact that P. magna is a recent invader in the Mediterranean (first
reported from Spain in 2000), it is recorded at various locations throughout the basin (Katsanevakis
et al. 2020a and references therein). P. magna is favoured under eutrophic conditions and forms
seasonally invasive populations that foul hard substrates in environments such as mussel farms,
lagoons and ports (Longo et al. 2007, 2012; Pierri et al. 2010; Guardiola et al. 2011; Halasz 2016;
Bachetarzi et al. 2019; Giangrande et al. 2020; Katsanevakis et al. 2020a). The invasive sponge
shows a particular preference to overgrow cultivated mussels, inducing potential negative effects
(i.e. affecting the mussel growth) on shellfish aquaculture production, and being a nuisance for
mussel farmers who make efforts to control the sponge growth (Longo et al. 2007, 2012;
Gerovasileiou et al. 2017). This fouling behavior has also been exhibited by the calcareous sponge
on top of the red algae Peyssonnelia squamaria (S.G. Gmelin) Decaisne ex J. Agardh, triggering
the activation of antioxidant defense mechanisms by the overgrown native red algae (Guzzetti et
al. 2019).
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Cnidaria
Macrorhynchia philippina Kirchenpauer, 1872
The feathery stinging hydroid Macrorhynchia philippina is an invasive species that has been
established in the eastern Mediterranean since the 1990’s (Bitar and Bitar-Kouli 1995; Bédry et al.
2021). It settles on hard substrates such as rocky reefs and artificial structures (Galil 2018). It is a
venomous species and contact with its polyps may cause pain, a burning and itching sensation,
lesions and weals that may last for days (Çinar et al. 2006; González-Duarte et al. 2016; Galil
2018; Bédry et al. 2021). An increase of M. philippiana density could negatively affect touristic
activities and local economies (Çinar et al. 2006; Uysal and Turan 2020; Bédry et al. 2021).
However, no impacts of the invasive hydroid have been reported so far on biodiversity or
ecosystem services.
Rhopilema nomadica Galil, Spanier & Ferguson, 1990
The nomad jellyfish Rhopilema nomadica, native to Indian Ocean, was firstly reported in the
Mediterranean in 1977 along the coast of Israel and has managed to spread in the eastern and
central Mediterranean (Galil et al. 1990; Lotan et al. 1994; Daly Yahia et al. 2013; Angel et al.
2016 and references therein; Balistreri et al. 2017; Edelist et al. 2020). Since the mid 1980s and
during every summer, the invasive jellyfish forms massive swarms, some extending over 100 km,
that often reach coastal waters and negatively affect fisheries, aesthetic value, tourism and
provisioning coastal facilities (Rilov and Galil 2009; Otero et al. 2013; Daly Yahia et al. 2013;
Galil and Goren 2014; Katsanevakis et al. 2014a; Angel et al. 2016 and references therein;
Madkour et al. 2019). In Israel, the annual loss of profit of beach visit restrictions due to R.
nomadica blooms has been estimated between 1.8–6.2 million € (Ghermandi et al. 2015). During
the latest years such blooms have started to regularly occur during the winter as well (Edelist et al.
2020). Furthermore, these massive blooms of the invasive jellyfish eventually lead to mass
mortalities of the invasive jellyfish that end up in the seabed and decompose (Guy-Haim et al.
2020). The decomposition of such enormous biomass quantities in the oligotrophic ecosystems of
the Levantine Sea leads to an environmental alteration with increased oxygen demand,
acidification and nutrient production (Guy-Haim et al. 2020). Microbial communities change and
ultimately may cause a decrease in primary production that will shift the already oligotrophic
ecosystems of the Levantine Sea to further degradation (Guy-Haim et al. 2020). R. nomadica
swarms are considered hazardous to human health as the invasive jellyfish contains venom in its
tentacles and has caused a great number of hospitalizations of swimmers and fishers in the eastern
Mediterranean due to its painful stings that may cause a variety of symptoms such as pain,
erythematous eruptions, swelling, itching and burning sensations, fever, chills, fatigue, muscular
aches and spasms, vomiting and even respiratory distress and anaphylactic reactions (Galil et al.
1990; Gusmani et al. 1997; Yoffe and Baruchin 2004; Öztürk and Isinibilir 2010; Çinar et al. 2011;
Friedel et al. 2016; Remigante et al. 2018; Galil 2018; Glatstein et al. 2018; Uysal and Turan 2020).
Furthermore, the invasive jellyfish is a voracious predator that consumes shrimps, mollusks and
fish larvae (Otero et al. 2013); the predation pressure of massive swarms can pose a threat for
native biodiversity (Galil and Goren 2014; Mannino et al. 2017a). R. nomadica predation on
plaktonic larvae has also been associated with veliger mortality and reduced recruitment of
24
Stramonita haemastoma (Linnaeus, 1767) (Rilov et al. 2001). However, Kuplik and Angel (2020)
state that R. nomadica should not be considered a major planktonic predator as it mainly feeds on
microscopic prey smaller than 150 μm (Kuplik and Angel 2020). It is possible that it competes for
these common food resources with organisms that possess a small mouth opening such as larval
fish stages (Kuplik and Angel, 2020). The establishment of the invasive jellyfish and the recurrent
massive blooms have caused changes in native fauna of the Levantine Sea, such as the replacement
of the other large-sized medusa Rhizostoma pulmo (Macri, 1778) (Galil 2000). Juveniles of the
commercially important Lessepsian fish Alepes djedaba (Forsskål, 1775) take shelter among R.
nomadica tentacles and thus the invasive jellyfish may have favoured the population increase of
the alien fish (Avian et al. 1995; Galil 2000). R. nomadica has been used as an effective
biomonitoring species of heavy metal pollution and radionuclides concentrations (Duysak et al.
2013; Mamish et al. 2015).
Oculina patagonica de Angelis, 1908
The cryptogenic scleractinian coral Oculina patagonica is established throughout the
Mediterranean Sea (Zenetos et al. 2010; Cutajar et al. 2020). It is an opportunistic benthic settler
that exhibits invasive behavior that is favoured by artificial substrates and antropogenic
disturbance (Salomidi et al. 2013; Serrano et al. 2013; García et al. 2015). It can form encrusting
colonies that negatively affect sessile communities by competition for space on the available hard
substrate (Sartoretto et al. 2008; Coma et al. 2011; Cutajar et al. 2020). While the cryptogenic
scleractinian exhibits a particular preference for artificial substrates in disturbed environments, it
also occurs in healthy ecosystems at natural hard substrates, mostly in the shallow infralittoral zone
(Çinar et al. 2006; Salomidi et al. 2013). O. patagonica colonies constitute a novel habitat, which
however diminishes structural complexity and species richness (Coma et al. 2011; Smith et al.
2014; Katsanevakis et al. 2014a). Among the impacted sessile species is the endangered
Mediterranean native colonial zooxanthellate scleractinian Cladocora caespitosa that is affected
by overgrowth by O. patagonica (Sartoretto et al. 2008). Well-developed erect macroalgal
communities appear to inhibit the proliferation of the cryptogenic scleractinian, while filamentous
algal turfs do not negatively affect it and may even enhance the coral’s settlement (Rubio-Portillo
et al. 2014). Still, the scleractinian is capable of negatively impacting native macroalgae and
causing a phase shift from complex macroalgal dominated communities to a prevalence of
encrusting coral bioconstructions (Serrano et al. 2012). Along with this shift, O. patagonica
impacts the provision of a variety of ecosystem services provided by macroalgae (Katsanevakis et
al. 2014a) such as food, biotic materials, cognitive benefits, recreation, symbolic and aesthetic
values, and life cycle maintenance (Salomidi et al. 2012). O. patagonica has been used as a model
species to study coral bleaching produced by Vibrio infection (Rubio-Portillo et al. 2020).
Ctenophora
Mnemiopsis leidyi A. Agassiz, 1865
25
Mnemiopsis leidyi is a ctenophore native to the coastal and estuarine waters of western Atlantic
that has been introduced throughout the Mediterranean Sea (Zenetos et al. 2010). In the Black Sea,
this invasive species has caused cascade effects through planktivory, with devastating
consequences on the food web and local fisheries (Katsanevakis et al. 2014a). In the Mediterranean
the species is found to develop seasonally abundant populations in lagoons and coastal ecosystems
(Galil 2012). This ctenophore is an opportunistic, voracious feeder that preys on small-sized
planktonic species of low swimming velocity, such as gastropod larvae and cirriped nauplii
(Marchessaux et al. 2020, 2021). Predation on zooplankton could favour eutrophication of coastal
ecosystems through the release of grazing pressure on phytoplankton (Marchessaux et al. 2021).
In addition to planktonic organisms, benthic taxa such as harpacticoid copepods and amphipods
are also included in the diet of M. leidyi, though in lower proportions (Marchessaux et al. 2021).
Although M. leidyi shares the same food resources with native jellyfish such as Aurelia spp., no
significant competitive impact has been detected so far (Marchessaux et al. 2021). On the contrary,
predation impacts have been locally observed in the Adriatic Sea, where the zooplankton
assemblage exhibited a reduction in biomass, species diversity and evenness soon after the
invasion of M. leidyi (Fiori et al. 2019). Massive blooms of this invasive ctenophore are reported
to negatively affect fisheries and other coastal activities due to the entanglement in nets and
clogging of water intake pipes (Galil 2012; Otero et al. 2013; Angel et al. 2016; Diciotti et al.
2016; Mytilineou et al. 2016).
Bryozoa
Amathia verticillata (delle Chiaje, 1822)
The spaghetti bryozoan Amathia verticillata was considered, until recently, to be of native origin
in the Mediterranean Sea. Eventually, A. verticillata was suggested to be a non-indigenous species
with an old presence across the Mediterranean basin (Galil and Gevili 2014), common and
widespread across the basin (Ferrario et al. 2018). In its global range, A. verticillata causes
significant negative impacts on biodiversity such as the loss of the canopies of Zostera marina
Linnaeus, 1753 due to overgrowth and shading by the erect bushy formations of its colonies
(McCann et al. 2015). It is considered a pest due to boat hull and fisheries equipment fouling and
the blocking of intake pipes on vessels and ponds (McCann et al. 2015). In the Mediterranean Sea,
A. verticillata is usually found inside marinas and ports, where it can form seasonally abundant
bushy colonies that extensively foul on facility structures (Marchini et al. 2015; Rizgalla et al.
2019a). Furthermore, A. verticillata colonies foul on aquaculture equipment and entangle in fishing
nets, causing economic losses for maintenance and cleaning purposes (Ounifi-Ben Amor et al.
2016; Yokeş et al. 2018). On the other hand, these dense colonial structures create a novel habitat
that supports a variety of macroinvertebrate species that settle within the bushy formations
(Marchini et al. 2015). A. verticillata possibly facilitates the spreading of some alien arthropod and
mollusk species that have been commonly found feeding and/or nestling among these colonies
(Dailianis et al. 2016; Furfaro et al. 2018). When abundant, A. verticillata constitutes an optimal
26
food source for the echinoid Paracentrotus lividus, one of the major grazers of the Mediterranean
(Camps-Castellà et al. 2020).
Tricellaria inopinata d'Hondt & Occhipinti Ambrogi, 1985
Tricellaria inopinata is an invasive bryozoan of Pacific origin that was first recorded in the
Mediterranean Sea in the Venice Lagoon, in 1982 (d’Hondt and Occhipinti-Ambrogi 1985;
Occhipinti-Ambrogi 2000). Ever since, the species has been reported from various Mediterranean
locations usually in harbors, marinas and canals mainly in Italy, but also in France, Tunisia, Greece
and Turkey (Ulman et al. 2017). T. inopinata is a fast-growing fouling species, capable of
outcompeting native bryozoan species and causing significant decline to their populations through
space competition and benthic community domination (Corriero et al. 2007; Occhipinti-Ambrogi
2000; Occhipinti-Ambrogi and Savini 2003). In Venice Lagoon this invasive bryozoan appears to
have caused dramatic changes in the structure of the bryozoan community and a significant decline
in populations of native bryozoans, in particular for species developing erect flexible colonies
(Occhipinti-Ambrogi 2000). T. inopinata has also been observed to foul the byssus thread and
form dense soft carpets on shells of the cultivated mussel Mytilus galloprovincialis, a settlement
with a potential negative impact on the production of mussel farms (Occhipinti-Ambrogi et al.
2000; Occhipinti-Ambrogi and Savini 2003; Fortič and Mavrič 2018; Fortič et al. 2019). Gavira-
O’Neill et al. (2018) also expressed concern about the ability of T. inopinata to compete with
native habitat-forming species and to promote the establishment of mobile alien species
(invasional meltdown sensu Simberloff and Von Holle, 1999) multiplying the potential impact of
former introduced species also at the expense of the native ones, a fact that could be devastating
for local biodiversity. Due to fouling on vessel hulls, T. inopinata causes the increase of fuel
consumption (Loxton et al. 2017).
Mollusca
Anadara kagoshimensis (Tokunaga, 1906)
The ark clam Anadara kagoshimensis, widely misidentified in the past literature from the
Mediterranean Sea as Anadara cornea (Reeve, 1844) or Anadara inaequivalvis (Bruguière 1789),
is native to the Indo-Pacific region and has invaded the Mediterranean during the 1960s (Zenetos
et al. 2004a). In the western Adriatic Sea, it has initially become a dominant species in soft bottom
habitats, impacting some native mollusks of commercial value such as the cardiid Cerastoderma
glaucum (Bruguière, 1789) (see Occhipinti-Ambrogi 2000; Streftaris and Zenetos 2006; Zenetos
et al. 2010). However, after an initial population explosion, it exhibited a strong regression phase
(Russo 2017; Strafella et al. 2017). Notwithstanding that, in the last decades new populations are
still being recorded in the Adriatic Sea (Despalatović et al. 2013a). As all filter-feeders, it filters
suspended particles and subsequently deposits faeces and unconsumed particles onto the
sediments, increasing sedimentation. Such increased sedimentation can represent a significant loss
of energy and nutrients from the water column and may decrease pelagic production (Katsanevakis
et al. 2014a). On the other hand, A. kagoshimensis is a bioturbator (ecosystem engineer), increasing
27
sediment water and oxygen content and releasing nutrients from the sediment to the water column
(Katsanevakis et al. 2014a; Smith et al. 2014). An increase in biomass and diversity of filter-
feeders following the invasion of A. kagoshimensis in Kazachya Bay (Crimea, Black Sea) led
Bondarev (2020) to argue in favour of possible positive effects on local biocenoses. As all
mollusks, it also absorbs carbonates to create its shell and thus has a potentially positive impact on
carbon sequestration and on climate regulation (Katsanevakis et al. 2014a). This role as a potential
sink of CO2 has still to be evaluated through an ecosystem approach — also accounting for
important ecosystem interactions, such as with phytoplankton populations and benthic-pelagic
coupling, which can significantly alter the CO2 cycle (Filgueira et al. 2019).
Anadara transversa (Say, 1822)
The ark clam Anadara transversa is native to the western Atlantic Ocean. It was first reported in
Mediterranean literature under its junior synonyms Anadara demiri (Piani 1981) and Scapharca
demiri Piani, 1981 (Demir 1977; Zenetos 1994). Subsequently, A. transversa spread and became
invasive in the Mediterranean (Zenetos et al. 2010), dominating soft-bottom benthic communities
and outcompeting native species, including species of commercial value such as the venerid
Chamelea gallina (Linnaeus, 1758) (see Morello et al. 2004). Juveniles can attach on aquaculture
ropes and may inhibit the settlement of mussels or compete with them for space (Morello et al.
2004; Dailianis et al. 2016; Nerlović et al. 2018). Adults burrow in the sediment, and thus they can
significantly increase pore-water flux and oxygen content of sediments and enhance nutrient
exchange between the sediments and the water column (Katsanevakis et al. 2014a; Smith et al.
2014). A. transversa is also: i) a structural engineer, as its empty shells may accumulate on the
seafloor providing shelter for other benthic organisms; ii) a bioturbator, modifying subtidal
sediments; iii) a filter-feeder, removing particles from the water column and thus reducing
turbidity and enhancing light penetration that favours a variety of organisms (Sousa et al. 2009;
Smith et al. 2014). As all filter-feeders, it filters suspended particles and subsequently deposits
faeces and unconsumed particles onto the sediments, increasing sedimentation. Such increased
sedimentation can represent a significant loss of energy and nutrients from the water column and
may decrease pelagic production (Katsanevakis et al. 2014a). As all mollusks, it absorbs
carbonates to create its shell and thus has a potentially positive impact on carbon sequestration and
on climate regulation (Katsanevakis et al. 2014a) This role as a potential sink of CO2 has still to
be evaluated through an ecosystem approach — also accounting for important ecosystem
interactions, such as with phytoplankton populations and benthic-pelagic coupling, which can
significantly alter the CO2 cycle (Filgueira et al. 2019).
Arcuatula senhousia (Benson, 1842)
The Asian bag mussel Arcuatula senhousia, native to the western Pacific Ocean, has invaded since
the 1960s various parts of the Mediterranean, where it mostly thrives inside sheltered bays and
estuaries (Crocetta et al. 2010; Despalatović et al. 2013b; Doğan et al. 2014; López-Soriano and
Quiñonero-Salgado 2018). In the soft substrates of such environments, it can form dense
aggregations and create byssal mats or even carpets, outcompeting native species and altering the
structure and functioning of invaded habitats (Mistri 2003; Otero et al. 2013; Katsanevakis et al.
2014a). At the same time, byssal mats and carpets created by A. senhousia act as novel structures
28
that enhance habitat heterogeneity, offering shelter to a variety of benthic organisms (Mistri 2003;
Mistri et al. 2003; Munari 2008; Katsanevakis et al. 2014a). No significant effects were detected
by A. senhousia mats on the growth of the clams Ruditapes decussatus and Ruditapes
philippinarum (A. Adams & Reeve, 1850) (Mistri 2004a). The species has also been observed to
overgrow benthic communities of hard substrates, and especially on the byssal threads of the
Mediterranean mussel Mytilus galloprovincialis at farmed populations (Cardone et al. 2014;
Crocetta et al. 2015a). A. senhousia is an effective phytoplankton filter-feeder that has been proven
to compete with other species and even reduce the available carbon transfer to benthic deposit
feeders, although only when occurring in high densities (Como et al. 2016, 2018). The rate at
which dense A. senhousia assemblages filter the water for food and oxygen accelerates the rate of
conversion of suspended sediment to deposited material and may rapidly alter the stability of the
substrate (Katsanevakis et al. 2014a). This is a process in which A. senhousia acts as a chemical
engineer, increasing organic deposited material onto sediments leading to the diminution of the
redox potential discontinuity layer, rendering the environment within or under byssal mats
unsuitable for adults or larvae of other species (Mistri et al. 2004). On the other hand, being a
filter-feeder that consumes suspended material from the water column, it reduces the turbidity and
increases the light penetration, thus improving environmental conditions of benthic photophilic
communities (Smith et al. 2014). As all mollusks, it absorbs carbonates to create its shell and thus
has a potentially positive impact on carbon sequestration and on climate regulation (Katsanevakis
et al. 2014a). Mistri and Munari (2013) assessed the balance among respiration, shell calcium
carbonate sequestration, and CO2 release during biogenic calcification and found that A. senhousia
is a net CO2 generator and thus the communities dominated by this invasive bivalve increase CO2
emissions from the sea to the atmosphere. This chemical approach has been subsequently criticized
as simplistic, ignoring important ecological interactions and mechanisms, such as with
phytoplankton populations and benthic-pelagic coupling; an ecosystem approach has been thus
suggested for assessing the role of bivalves as a potential sink of CO2 (Filgueira et al. 2019).
Brachidontes pharaonis (P. Fischer, 1870)
The rayed Erythrean mussel Brachidontes pharaonis is an alien mytilid of Indo-Pacific origin
introduced in the Mediterranean since more than a century; it is currently established with
abundant populations in the eastern and central parts of the basin (Zenetos et al. 2004a; Bonnici et
al. 2012; Crocetta et al. 2013), extending also to the Adriatic (Lipej et al. 2017) and the western
Mediterranean (Di Natale 1982). Studies conducted at the Levantine coasts of Israel during the
1970s originally showed no competitive effect of the alien bivalve against the small Mediterranean
mussel Mytilaster minimus (Poli, 1795) (see Galil 2007a). However, by late 1990s, B. pharaonis
exhibited a dominance shift forming dense aggregations of up to 300 specimens per 100 cm²
(Mienis 2003; Rilov et al. 2004; Galil 2007a). Its biological traits then allowed it to outcompete or
even almost totally replace native bivalves in rocky benthic communities at several sites in the
eastern and even central Mediterranean Sea, impacting the native vermetid platforms and creating
novel habitats that facilitated the arrival of additional alien species but also providing shelter for
native ones (Safriel and Sasson-Frostig 1988; Galil 2007a; Rilov and Galil 2009; Crocetta et al.
2013; Katsanevakis et al. 2014a; Smith et al. 2014; Çinar et al. 2017). Recently, easternmost
Mediterranean populations substantially declined in number (Rilov et al. 2020), although the
29
numbers now slowly recover. Notwithstanding such premises and the habitat-forming traits
described above, a comparative study conducted in 2009 on an invasive B. pharaonis population
in Malta revealed lower values of species richness in comparison to an adjacent rocky reef still
characterized by the absence of the rayed Erythrean mussel (Bonnici et al. 2012). As all filter-
feeders, it filters suspended particles and subsequently deposits faeces and unconsumed particles
onto the sediments, increasing sedimentation. Such increased sedimentation can represent a
significant loss of energy and nutrients from the water column and may decrease pelagic
production (Katsanevakis et al. 2014a). Still, the species can be a crucial component of ecosystems
that reduces water turbidity and contributes to the community by supporting other consumers
through its filter-feeding abilities and the formation of novel habitats (Sarà et al. 2021). As all
mollusks, it absorbs carbonates to create its shell and thus has a potentially positive impact on
carbon sequestration and on climate regulation (Katsanevakis et al. 2014a). This role as a potential
sink of CO2 has still to be evaluated through an ecosystem approach — also accounting for
important ecosystem interactions, such as with phytoplankton populations and benthic-pelagic
coupling, which can significantly alter the CO2 cycle (Filgueira et al. 2019). The rayed Erythrean
mussel may also induce substantial costs by fouling intake pipes, heat exchangers, and underwater
constructions (Garaventa et al. 2012; Otero et al. 2013). Finally, the native whelk Stramonita
haemastoma has preferably incorporated B. pharaonis in its diet, due to the high energy gain it
provides as a food source (Rilov et al. 2002; Giacoletti et al. 2016, 2017).
Bursatella leachii Blainville, 1817
The ragged sea hare Bursatella leachii is a circumtropical species widespread in the Atlantic and
the Indo-Pacific regions. Early Mediterranean records from the Levantine Sea led researchers to
consider it a Lessepsian immigrant for almost a century (see reviews in Selfati et al. 2017;
Travaglini and Crocetta 2019). However, recent molecular work clarified its dispersal pathway,
suggesting that Mediterranean populations originated from the Atlantic (Bazzicalupo et al. 2018,
2020). This led to debates about its alien status, with some researchers considering it as a crypto-
expanding species or neonative, excluding it from national lists (e.g. Servello et al. 2019; Crocetta
et al. 2020) and others still listing it as an alien species (Petović et al. 2019; Katsanevakis et al.
2020b; Zenetos and Galanidi 2020; Zenetos et al. 2020a). Notwithstanding such disputes, B.
leachii is indeed one of the most successful recent colonizers of the Mediterranean Sea, with
records from 19 out of 23 Mediterranean countries (Rizgalla and Crocetta 2020). B. leachii forms
seasonally abundant populations in shallow sediments of degraded and eutrophic environments
and has been also observed clogging in fishing nets inside lagoons (Lipej et al. 2012; Selfati et al.
2017; González-Wangüemert et al. 2018), although no negative impacts on biodiversity have been
ever reported. Sulfated polysaccharides extracted from B. leachii exhibited anticoagulant activity,
thus suggesting that extracts from this mollusk may be a promising alternative in anticoagulant
therapies (Dhahri et al. 2020), and, in general, bioactive compounds from this species also showed
antioxidant, neuroprotective, and anti-inflammatory activities (Braga 2014).
Chama pacifica Broderip, 1835
The Erythrean jewel box oyster Chama pacifica is native of the Indo-Pacific and widespread in
the eastern Mediterranean basin, where it is found cemented to massive rocks in exposed areas
30
from the midlittoral to infralittoral zone (Zenetos et al. 2004a; Crocetta and Russo 2013). Its
population sizes strongly increased during the last decades, replacing a number of native habitat
formers (Galil 2007a; Crocetta et al. 2013a; Otero et al. 2013). These dense populations can form
reefs and beds, completely changing the structure of the invaded habitats and the composition of
their benthic communities (Crocetta et al. 2013a; Otero et al. 2013). The Erythrean jewel box
oyster also competes with other filter-feeders for the available food resources (Otero et al. 2013).
On the other hand, C. pacifica reefs provide services similar to oyster beds, such as water quality
regulation and bioremediation of waste by biofiltration, and support reproduction of a variety of
species by acting as nursery areas (Katsanevakis et al. 2014a). In fact, they enhance spatial
complexity and offer an appropriate habitat to a variety of benthic organisms; even the empty
shells and loose valves of C. pacifica offer a fitting habitat for a diverse fouling community of
algae and invertebrates (Fishelson 2000; Streftaris and Zenetos 2006; Katsanevakis et al. 2014a;
Shabtay et al. 2014). As all filter-feeders, it filters suspended particles and subsequently deposits
faeces and unconsumed particles onto the sediments, increasing sedimentation. Such increased
sedimentation can represent a significant loss of energy and nutrients from the water column and
may decrease pelagic production (Katsanevakis et al. 2014a). As all mollusks, it absorbs
carbonates to create its shell and thus has a potentially positive impact on carbon sequestration and
on climate regulation (Katsanevakis et al. 2014a). This role as a potential sink of CO2 has still to
be evaluated through an ecosystem approach — also accounting for important ecosystem
interactions, such as with phytoplankton populations and benthic-pelagic coupling, which can
significantly alter the CO2 cycle (Filgueira et al. 2019). C. pacifica has been finally assessed as a
potentially promising food source for market exploitation in comparison to native bivalves of
Turkey (Tanrıverdi et al. 2020), although this species may accumulate heavy metals in edible parts
(Türkmen et al. 2005).
Conomurex persicus (Swainson, 1821)
The Persian conch Conomurex persicus is a species that originates in the Persian Gulf and has
invaded the Mediterranean since 1978 (Nicolay 1986). In a few decades, it has become
overabundant in the sandy habitats of the eastern Mediterranean, where it has exhibited invasive
behaviour (Mienis 2004; Guarnieri et al. 2017; Crocetta et al. 2017, 2020; Zenetos et al. 2018). Its
presence in the eastern Mediterranean has been associated with a depletion of the rhodophyte Jania
rubens (Linnaeus) J.V. Lamouroux due to potential grazing (Mutlu 2004). On the other hand,
empty shells of C. persicus are used by Mediterranean hermit crabs and thus can positively affect
hermit crab populations, in particular in shell-limited cases (Mutlu and Ergev 2010; Özcan et al.
2013). C. persicus is exploited and marketed for human consumption in various localities of the
Mediterranean (Mienis 1999; Rilov and Galil 2009; Katsanevakis et al. 2008; Pancucci-
Papadopoulou et al. 2012; Corsini-Foka et al. 2015; Crocetta et al. 2017). In addition, it has been
a subject for biological experiments, and a novel 40 kD protein (ACLS40) has been extracted from
the gastropod’s shell, with the ability to stabilize calcium carbonate in the form of the
thermodynamically unstable vaterite polymorph (Pokroy et al. 2006). Finally, C. persicus can be
also used effectively as a biomonitoring species for radionuclides detection (Ammar et al. 2009).
As all mollusks, it absorbs carbonates to create its shell and thus has a potentially positive impact
on carbon sequestration and on climate regulation (Katsanevakis et al. 2014a). This role as a
31
potential sink of CO2 has still to be evaluated through an ecosystem approach — also accounting
for important ecosystem interactions, which can significantly alter the CO2 cycle (Filgueira et al.
2019).
Crepidula fornicata (Linnaeus, 1758)
The American slipper limpet Crepidula fornicata is native to the Atlantic coast of the USA and
invasive in Europe, including the Mediterranean Sea (e.g. Macali et al. 2013; Dragičević et al.
2019; Zenetos et al. 2020b). The species is a pest in some localities of the Atlantic Ocean (see
Otero et al. 2013; Katsanevakis et al. 2014a), but population densities of this species are low in the
majority of the sites of occurrence in the Mediterranean Sea, and thus no studies aimed to assess
its local impacts in the basin. As all filter-feeders, it filters suspended particles and subsequently
deposits faeces and unconsumed particles onto the sediments, increasing sedimentation. Such
increased sedimentation can represent a significant loss of energy and nutrients from the water
column and may decrease pelagic production (Katsanevakis et al., 2014). As all mollusks, it
absorbs carbonates to create its shell and thus has a potentially positive impact on carbon
sequestration and on climate regulation (Katsanevakis et al., 2014). This role as a potential sink of
CO2 has still to be evaluated through an ecosystem approach — also accounting for important
ecosystem interactions, which can significantly alter the CO2 cycle (Filgueira et al. 2019).
Dendostrea cf. folium (Linnaeus, 1758)
Molecular identification of Mediterranean specimens recently ascribed to Dendostrea folium
(Linnaeus, 1758) [and misidentified in the past as Dendostrea frons (Linnaeus, 1758)] turned out
to be problematic, as distance values of Mediterranean vs. Indo-Pacific samples support a close
relationship, yet are not conclusive for testing conspecificity (Crocetta et al. 2015b). Therefore,
we report here the Mediterranean taxon as Dendostrea cf. folium. The Dendostrea taxon that
invaded the Mediterranean basin originates in the Indo-Pacific, and in a few decades it spread in
the entire eastern Mediterranean Sea, with sporadic records from the central parts of the basin and
the Adriatic Sea (Çeviker 2001; Crocetta et al. 2013, 2017; Karachle et al. 2016; Gerovasileiou et
al. 2017; Ulman et al. 2017). The species is now very common in shallow waters of the Levantine
Sea, where it creates aggregations and small reefs, and it also fouls aquaculture fish cages and
mussel farms (F. Crocetta, unpublished data). It may potentially enhance the complexity of the
substrates, thus supporting the arrival of new settlers, but may also reduce water flow and oxygen
levels, thus causing economic losses. As all filter feeders, it filters suspended particles and
subsequently deposits faeces and unconsumed particles onto the sediments, increasing
sedimentation. Such increased sedimentation can represent a significant loss of energy and
nutrients from the water column and may decrease pelagic production (Katsanevakis et al. 2014a).
As all mollusks, it absorbs carbonates to create its shell and thus has a potentially positive impact
on carbon sequestration and on climate regulation (Katsanevakis et al. 2014a). This role as a
potential sink of CO2 has still to be evaluated through an ecosystem approach — also accounting
for important ecosystem interactions, which can significantly alter the CO2 cycle (Filgueira et al.
2019). Noteworthy, no studies aimed to assess its impacts on biodiversity or ecosystem services
in Mediterranean ecoregions.
32
Fulvia fragilis (Forsskål in Niebuhr, 1775)
The fragile cockle Fulvia fragilis, often misidentified in the past literature from the Mediterranean
Sea as Fulvia papyracea (Bruguière, 1789), is native of the Indo-Pacific region and well
established in the Mediterranean, with multiple records spanning from the eastern to the western
Mediterranean Sea (Crocetta 2005; Goud and Mifsud 2009; López-Soriano et al. 2009; Crocetta
et al. 2013; Rizgalla et al. 2019b; Zenetos et al. 2020a) including the Adriatic (Gerovasileiou et al.
2017; Gvozdenović et al. 2019). Despite of that, studies on the species have been only carried out
along the Mediterranean coast of Africa (Mahmoud et al. 2010; Rifi et al. 2011, 2012, 2013, 2015a,
2015b), which mostly revealed that the species is a simultaneous hermaphrodite, with a continuous
spawning all along the year, and that it constitutes a valid biomonitoring indicator of aquatic
pollution. With regards to its competition with native species, Ben Souissi et al. (2003) initially
stated that the spread of this species was strongly limited by the competition with other native
bivalve species such as Acanthocardia paucicostata (G. B. Sowerby II, 1834) and Cerastoderma
glaucum (Bruguière, 1789). However, the fragile cockle spread in all Tunisian lagoons after a few
decades, outcompeting native bivalves and even causing dreadful declines to their populations
(Ounifi-Ben Amor et al. 2016). As all filter feeders, it filters suspended particles and subsequently
deposits faeces and unconsumed particles onto the sediments, increasing sedimentation. Such
increased sedimentation can represent a significant loss of energy and nutrients from the water
column and may decrease pelagic production (Katsanevakis et al. 2014a). As all mollusks, it
absorbs carbonates to create its shell and thus has a potentially positive impact on carbon
sequestration and on climate regulation (Katsanevakis et al. 2014a). This role as a potential sink
of CO2 has still to be evaluated through an ecosystem approach — also accounting for important
ecosystem interactions, which can significantly alter the CO2 cycle (Filgueira et al. 2019). Various
studies in the Mediterranean have assessed this invasive bivalve as a valid biomonitoring indicator
of aquatic pollution (Mahmoud et al. 2010; Rifi et al. 2015a).
Magallana/Crassostrea species
The taxonomic status of Magallana species, and mainly of Magallana gigas (Thunberg, 1793) and
Magallana angulata (Lamarck, 1819) [formerly belonging to the genus Crassostrea: see Salvi and
Mariottini 2020], is still controversial, with some authors considering them as different species
(e.g. Lapègue et al. 2004; Liu et al. 2011; Hsiao et al. 2016) and others regarding them as
conspecific (e.g. López-Flores et al. 2004; Reece et al. 2008). They are here conservatively treated
as different sibling species. Moreover, almost all records of Magallana specimens in the
Mediterranean Sea were based on shell morphology, thus suggesting that records of both species
were often mixed up in the past literature from the Mediterranean Sea under the binomial name
“Crassostrea gigas”. One of the exceptions is constituted by the paper by Fabioux et al. (2002),
who recorded the presence of both M. gigas and M. angulata in the Adriatic Sea based on
molecular data. To further complicate matters, another alien oyster species,Crassostrea virginica
(Gmelin, 1791), has been also reported from the Mediterranean Sea based on shell morphology
only (e.g. Cossignani et al. 1992; Goigne 2001). Thus, in absence of any sort of certainty, these
alien oysters are here reported as “Magallana/Crassostrea species”. Species of this group were
intentionally introduced in various Mediterranean countries by the aquaculture industry since at
least the beginning of the 19th Century (Monterosato 1915; Dantan and Heldt 1932; Mazzarelli
1936), and are currently still cultivated with important revenues (Zenetos et al. 2010; Otero et al.
33
2013; Katsanevakis et al. 2014a; Galil et al. 2018). Noteworthy, the oyster farming production is
recently suffering of important losses worldwide due to a virus disease (Ostreid herpesvirus 1,
OsHV-1) that is causing massive mortalities especially during summer periods (Bookelaar 2018;
Ugalde et al. 2018). Mass mortality events of the species in lagoons can alter environmental
conditions through NH4 excretion, PO4 concentration increase and N:P ratio decrease due to oyster
tissue decomposition, and eventually lead to a strong modification of the planktonic microbial
community structure and an increase of picoplantkon and ciliates (Richard et al. 2019). Alien
oysters have also escaped from aquaculture facilities in the past and created wide “natural” banks
and oyster reefs. These banks completely alter habitat structure and often dominate even the
infralittoral fringe of the invaded areas, competing for space and resources with native mussels
and oysters and determining negative effects (Crocetta 2011; Otero et al. 2013; Katsanevakis et al.
2014a; Ezgeta-Balić et al. 2020) on provided ecosystem services (Salomidi et al. 2012). On the
other hand, these novel habitats support a new variety of species (Streftaris and Zenetos 2006;
Katsanevakis et al. 2014a; Smith et al. 2014), contributing to potentially enhancement of
biodiversity even when dead because loose valves offer a fitting habitat for a diverse fouling
community of algae and invertebrates. The newly formed reefs contribute significantly to coastal
protection as they stabilize soft substrates as long as they persist (Katsanevakis et al. 2014a). Still
the finding of larvae of Ostrea edulis Linnaeus, 1758 in stomachs of alien oysters also revealed a
predator-prey relationship with a potential negative effect on the vulnerable wild populations of
the native oyster in the Adriatic (Ezgeta-Balić et al. 2020). Goedknegt et al. (2020) demonstrated
that the habitat structure provided by invasive Pacific oysters can also indirectly affect parasitism
in native blue mussels. Magallana/Crassostrea species are also valid biomonitoring taxa for
tracing the introduction or persistence of harmful elements (i.e. Cd, Cu, Zn) in aquatic
environments due to accumulation in their tissues (Burioli et al. 2017). Increased biofiltration due
to the presence of wide oyster banks removes sedimentary material and phytoplankton from the
water column, thus increasing water clarity and creating the appropriate light conditions for
vegetation to grow in deeper waters (Deslous-Paoli et al. 1998). At the same time, these species
filter suspended particles and subsequently deposit faeces and unconsumed particles onto the
sediments, increasing sedimentation. Such increased sedimentation can represent a significant loss
of energy and nutrients from the water column and may decrease pelagic production (Katsanevakis
et al. 2014a). As all mollusks, alien oysters absorb carbonates to create their shells and thus have
a potentially positive impact on carbon sequestration and on climate regulation (Katsanevakis et
al. 2014a), although this role as a potential sink of CO2 has still to be evaluated through an
ecosystem approach — also accounting for important ecosystem interactions, which can
significantly alter the CO2 cycle (Filgueira et al. 2019). Socioeconomic impacts include injuries
caused by shells on leisure beaches and blockage of navigational channels for recreational vessels
(Herbert et al. 2016).
Mya arenaria Linnaeus, 1758
The soft-shell clam Mya arenaria is native to the Atlantic coast of the USA and invasive in Europe,
including records from the Mediterranean Sea (Zenetos et al. 2010; Crocetta and Turolla 2011;
Zenetos et al. 2020b). In the Black Sea, the Atlantic Ocean, and the Baltic Sea, M. arenaria is
completely naturalized and this makes it difficult to identify its impact. Yet, M. arenaria has been
reported to cause significant impacts on the Black Sea ecosystem health and fish production
(Kasapoglu et al. 2011). Population densities of the soft-shell clam are low in the majority of the
34
sites of occurrence in the Mediterranean Sea, and thus no studies aimed to assess its local impacts
in the basin.
Petricolaria pholadiformis (Lamarck, 1818)
The American piddock Petricolaria pholadiformis is native to the Atlantic coast of the USA and
invasive in Europe, including confirmed records from the Mediterranean Sea (Saronikos Gulf,
Greece) (Zenetos et al. 2009a; Crocetta et al. 2017; Zenetos et al. 2020b). As its population density
in the Mediterranean Sea is very low, no studies have to date assessed its local impacts in the basin.
Pinctada radiata (Leach, 1814)
Specimens morphologically ascribed to the genus Pinctada Röding, 1798 belong to a complex of
three related species: Pinctada imbricata (Röding, 1798) distributed in the western Atlantic region,
Pinctada radiata (Leach, 1814) distributed in the eastern Indian Ocean and the Red Sea regions,
and Pinctada fucata (A. Gould, 1850) distributed in the Indo-Pacific region. The taxonomy of
these congeneric cryptic species was molecularly investigated by two articles (Tëmkin 2010;
Cunha et al. 2011), published about simultaneously. Tëmkin (2010) treated P. fucata and P.
radiata as geographical subspecies of P. imbricata, whereas Cunha et al. (2011) presented a
molecular tree that resolved them at the rank of species. The pearl oyster is one of the first
Mediterranean invaders after the opening of the Suez Canal, and has now spread all over the
Mediterranean Sea, often reaching very high densities (Zenetos et al. 2004b; Crocetta et al. 2009,
2013, 2017; Gerovasileiou et al. 2017; Ballesteros et al. 2020). When abundant, it can form reefs
that completely alter habitat structure and impact native fauna, outcompeting native filter-feeding
species for food resources and space (Otero et al. 2013; Katsanevakis et al. 2014a). It has also been
found to massively overgrow the rhizomes of Posidonia oceanica, possibly affecting the fitness of
the meadow (Ounifi-Ben Amor et al. 2016). At the same time, reefs also contribute to enhance
habitat complexity and native biodiversity by offering substrate and shelter for other settlers, and
provide a variety of ecosystem services such as food provision, water purification by biofiltration,
and life cycle maintenance for a variety of species (Katsanevakis et al. 2014a). Pinctada radiata
is known to create dense fouling aggregations on aquaculture fish cages and mussel farms that
reduce water flow and oxygen levels within the cultivation units, causing economic losses (Otero
et al. 2013; Deidun et al. 2014; Katsanevakis et al. 2014a; Ounifi-Ben Amor et al. 2016; Theodorou
et al. 2019). It is also a tolerant species to chemical pollution and has been used numerous times
as a biomonitoring species for tracing heavy metals in the marine environment (Zenetos et al.
2004b; Göksu et al. 2005; Sakellari et al. 2013; Banana et al. 2016). Lipids extracted from the
nacre and flesh stimulate human bone formation and could be potentially used as a medicine
against osteoporosis (Ben Ammar et al. 2019). Furthermore, protein extracted from the species can
have antibacterial effects, making the invasive bivalve a potential source for pharmaceutical
products (Mona et al. 2018). Finally, P. radiata is an edible bivalve (Ben Ammar et al. 2014), and
cultivation attempts have been also carried out in various countries of the Mediterranean
(Katsanevakis et al. 2008). As all filter feeders, it filters suspended particles and subsequently
deposits faeces and unconsumed particles onto the sediments, increasing sedimentation. Such
increased sedimentation can represent a significant loss of energy and nutrients from the water
column and may decrease pelagic production (Katsanevakis et al. 2014a). As all mollusks, it
35
absorbs carbonates to create its shell and thus has a potentially positive impact on carbon
sequestration and on climate regulation (Katsanevakis et al. 2014a), although this role as a
potential sink of CO2 has still to be evaluated through an ecosystem approach — also accounting
for important ecosystem interactions, which can significantly alter the CO2 cycle (Filgueira et al.
2019).
Rapana venosa (Valenciennes, 1846)
The Asian whelk Rapana venosa is native to the Sea of Japan, Yellow Sea and East China Sea and
alien in the Mediterranean Sea, where it was recorded in the north Adriatic Sea in 1973 (Ghisotti
1974). R. venosa is a voracious predator with a preference for mussels, oysters and other bivalves.
The species has also invaded the Black Sea and caused significant predation impacts on native
mollusks. In 2006 and 2007, massive settlement of Rapana juveniles occurred at many sites along
the Bulgarian coasts, resulting in a complete coverage of the substrate with a compact blanket of
Rapana. The bivalve populations of the area were completely obliterated, even the smallest
species, such as Mytilaster lineatus (Gmelin, 1791). Its predation impact has also been evident in
Ukranean waters, where it has destroyed oyster banks (Katsanevakis et al. 2014a). Such negative
impacts on mussel and oyster beds and reefs, negatively affect the provision of various ecosystem
services (Salomidi et al. 2012; Katsanevakis et al. 2014a; Zenetos and Galanidi 2017). On the other
hand, the species is edible and has been exploited by the Black Sea fisheries (Katsanevakis et al.
2014a and references therein), yielding an important revenue throughout the years especially for
Turkey and Bulgaria (Zenetos and Galanidi 2017). In the Mediterranean Sea, the species is known
from a wide number of separate records (e.g. Terreni 1980; Cesari and Pellizzato 1985;
Koutsoubas and Voultsiadou-Koukoura 1991; Crocetta and Soppelsa 2006; see Zenetos and
Galanidi 2017 for a full account). Occasionally Rapana whelks can be seen in fish markets of the
northern Adriatic (Clodia database, 2020). Food provisioning services were impacted in the past
in Italy, where R. venosa disrupted the squid fishery by increasing the weight of the nets used for
this type of fishing that are more easily damaged (Savini et al. 2004). Despite that and after an
initial expansion, R. venosa exhibited a strong regression phase, and now it does not raise a concern
for local fisheries or native bivalve populations (F. Crocetta, unpublished data; Savini and
Occhipinti-Ambrogi 2006; Occhipinti-Ambrogi et al. 2011). Its operculum has traditionally been
used in Chinese medicine and was also used as an ingredient for incense production during ancient
and medieval years (Nongmaithem et al. 2017).
Ruditapes philippinarum (A. Adams & Reeve, 1850)
The manila clam Ruditapes philippinarum is an alien bivalve of Indo-Pacific origin, considered as
one of the most productive aquatic farming species worldwide (Astorga 2014). It was first
introduced in the Mediterranean Sea (in Italy and France) by the aquaculture industry to
compensate for the irregular yields of the native congeneric species Ruditapes decussatus, and in
a few years, it has become an important biological resource, significantly contributing to
aquaculture and fishery sectors (Otero et al. 2013; Katsanevakis et al. 2014a). A lifecycle
assessment (LCA) in the Adriatic Sea also showed R. philippinarum farming to be a sustainable
practice of aquaculture, at least in comparison to other native and alien mussels and oysters
(Turolla et al. 2020). The cultivated clams capture large amounts of carbon dioxide, transforming
36
cultivation areas to carbon sinks and thus mitigating anthropogenic pollution impacts
(Katsanevakis et al. 2014a; Turolla et al. 2020). However, this role as a potential sink of CO2 has
still to be evaluated through an ecosystem approach, also accounting for important ecosystem
interactions, which can significantly alter the CO2 cycle (Filgueira et al. 2019). Furthermore, an
important amount of nutrients is also stored in clam shells, contributing to eutrophication reduction
(Turolla et al. 2020). R. philippinarum is an effective bioturbator that restructures surface
sediments, increasing sediment erosion and resuspension rates and reducing sediment stability
(Sgro et al. 2005; Smith et al. 2014). R. phillipionarum has exhibited significant bioturbation
activity that influences the distribution of oxygen in sediments, enhances solute exchange with the
overlying water column, and affects nutrient cycling (Bartoli et al. 2001; de MouraQueirós et al.
2011). Dense R. philippinarum populations can increase oxygen and carbon dioxide fluxes,
accelerate nutrient circulation and promote phytoplankton blooms and macroalgal growth (Bartoli
et al. 2001). The rapid circulation of nutrients due to R. philippinarum dense populations can
promote new phytoplankton blooms and also sustain macroalgal growth, and thus positively affect
primary production (Bartoli et al. 2001). The manila clam is a particularly successful invader of
the Mediterranean and as an effective filter-feeder it competes for food with various native species
and even displaces some of them, including the native R. decussatus (Pranovi et al. 2006; Zenetos
et al. 2010; Otero et al. 2013). The two congeneric species also hybridize, a process that is probably
facilitated by their seasonal spawning overlap (Hurtado et al. 2011). Noteworthy that, the manila
clam introduction provides an additional food source for shellfish-eaters, such as shorebirds, in
southern England (Caldow et al. 2007). In the Venice lagoon, it was estimated that the introduction
of R. philippinarum doubled the benthos’ filtration capacity, which altered the functioning of the
ecosystem, resulting in a stronger benthic–pelagic coupling (Pranovi et al. 2006). The last authors
also stated that, after the introduction of R. philippinarum, “the Venice Lagoon ecosystem has
entered into a new state, probably more resistant but less resilient”. Manila clam farms have
operated as a vector for the introduction of additional alien species, invertebrates and algae that
attach to packaging material, foul R. philippinarum shells, or parasitize its tissues (Savini et al.
2010; Boudouresque et al. 2020). R. philippinarum has exhibited a prominent response when
acting as a bioindicator species to monitor elevated heavy metal concentrations, hydrocarbon
pollution, or the presence of polybrominated diphenyl ethers (PBDEs) in the water column (Pizzini
et al. 2015; Aru et al. 2016; Cacciatore et al. 2018).
Spondylus spinosus Schreibers, 1793
The spiny oyster Spondylus spinosus is native of the Indo-Pacific region and well established in
the entire eastern Mediterranean Sea, where it was first recorded in 1987–1988 at the Israeli coasts
(Zenetos et al. 2004a, 2009b; Delongueville and Scaillet 2006; Crocetta et al. 2013). Within just a
few decades, it thrived in the local rocky communities, creating dense reefs, altering previous
habitat features, and outcompeting or replacing native filter-feeding bivalves, including the
congeneric Spondylus gaederopus Linnaeus, 1758 (see Mienis et al. 1993; Crocetta et al. 2013);
moreover, such crowded reefs significantly deplete plankton availability. At the same time, it also
contributed to enhance habitat complexity and native biodiversity by offering substrate and shelter
for other settlers (Fishelson 2000; Delongueville and Scaillet 2006; Streftaris and Zenetos 2006;
Crocetta et al. 2013; Otero et al. 2013; Shabtay et al. 2014; Smith et al. 2014;). S. spinosus is
37
exploited and marketed for human consumption in various localities of the Mediterranean (Rilov
and Galil 2009; van Gemert 2017). It has also been successfully assessed as a bioindicator for the
presence of toxic elements in the marine environment (Ghosn et al. 2020; Shoham-Frider et al.
2020). Finally, high concentrations of microplastics have been also determined in soft tissues of
Levantine populations (Kazour 2019). All this makes this species, under certain circumstances, a
potential health concern when consumed. As all filter feeders, it filters suspended particles and
subsequently deposits faeces and unconsumed particles onto the sediments, increasing
sedimentation. Such increased sedimentation can represent a significant loss of energy and
nutrients from the water column and may decrease pelagic production (Katsanevakis et al. 2014a;
Shabtay et al. 2014). As all mollusks, it absorbs carbonates to create its shell and thus has a
potentially positive impact on carbon sequestration and on climate regulation (Katsanevakis et al.
2014a), although this role as a potential sink of CO2 has still to be evaluated through an ecosystem
approach — also accounting for important ecosystem interactions, which can significantly alter
the CO2 cycle (Filgueira et al. 2019).
Annelida:Polychaeta
Ficopomatus enigmaticus (Fauvel, 1923)
The serpulid Ficopomatus enigmaticus is an alien reef-building polychaete, established throughout
the Mediterranean (Zenetos et al. 2010). It represents a species complex composed by at least three
cryptic or pseudocryptic species (Styan et al. 2017); molecular data outside of the native range are
still scanty, but more than one species seem to occur in the Mediterranean Sea (Grosse et al. 2021).
This tubeworm is an ecosystem engineer that creates reef-like sessile colonies with vertically
developed serpulid tubes on hard substrate that attach to each other in clumps (Schwindt et al.
2001; Katsanevakis et al. 2014a). The newly formed biogenic reefs provide habitat and shelter for
many benthic epibionts (Bianchi and Morri 1996; Katsanevakis et al. 2014a; Smith et al. 2014).
On the other hand, previous space occupiers lose their habitat and are being displaced by the novel
formations (Katsanevakis et al. 2014a). In the Mediterranean such dense aggregations of this
invasive polychaete that forms biogenic reefs are mainly found inside transitional or eutrophic
ecosystems (Bianchi and Morri 1996, 2001; Fornós et al. 1997; Despalatović et al. 2013b; Cardone
et al. 2014; Saint Martin and Saint Martin 2015; Langeneck et al. 2015; Longo et al. 2016;
Tempesti et al. 2020). F. enigmaticus is an effective filter-feeder with the ability to consume large
amounts of suspended particles and thus to reduce water turbidity, phytoplankton concentration,
and large amounts of inorganic nutrients (Davies et al. 1989; Bruschetti et al. 2008; Katsanevakis
et al. 2014a). Nevertheless, such alterations in invaded lagoons can have dramatic effects on native
biodiversity (Katsanevakis et al. 2014a). F. enigmaticus has the potential to negatively impact
ecosystem services such as food provision, water storage and provision, symbolic and aesthetic
values and recreation and tourism due to fouling on water intake pipes, fishing boats, leisure craft,
floating structures in lagoons and docks, aquaculture ponds, ports, and docks (Katsanevakis et al.
2014a). The species has also shown potential to be used as an indicator species for ecotoxicological
assessment (De Marchi et al. 2019; Oliva et al. 2019, 2020; Cuccaro et al. 2021).
38
Hydroides elegans (Haswell, 1883)
The serpulid polychaete Hydroides elegans was one of the first recorded alien species of the
Mediterranean, reported from Naples in 1870 (Galil and Goren 2014) and potentially occurring
since the first half of the 19th Century (Langeneck et al. 2020). It is currently established in most
countries throughout the basin (Zenetos et al. 2010; Ulman et al. 2017). Molecular data suggest
that H. elegans represents in fact a species complex, with at least two cryptic species occurring in
the Mediterranean Sea (Grosse et al. 2021). H. elegans is an invasive serpulid that fouls on hard
substrates, mainly on eutrophic environments and artificial surfaces, creating dense aggregations
of calcified tubes (Antoniadou et al 2011; Sandonnini et al. 2021a, b) that constitute a new habitat
for new settlers (Koçak et al. 1999; Çinar et al. 2008). H. elegans dominates benthic communities
and competes with native species for space and resources (Elsayed and Dorgham 2019; Mangano
et al. 2019; Fortič et al. 2021). H. elegans also competes with its native congeneric Hydroides
dianthus (Verril, 1873) (Bianchi 1983). It fouls aquaculture facilities, causing economic losses
mainly due to maintenance and cleaning procedures (Relini 1993; Antoniadou et al. 2013;
Giangrande et al. 2020). Its fouling formations become a nuisance for many other artificial
structures such as quays, ship hulls and clog cooling systems (Katsanevakis et al. 2014a). In
contrast with other gregarious serpulids, growth is extremely fast, with a span of three weeks
between settlement and spawning (Nedved and Hadfield 2009). Although spawning in subtropical
and tropical areas occurs all year long, with a possible reduction in summer due to hypoxic
conditions and high salinity (Qiu and Qian 1998; Shin et al. 2013), in the Mediterranean Sea this
species seems to increase in spring and densely cover docks and ship hulls in summer (Lezzi et al.
2018; Mangano et al. 2019), in correspondence with the touristic season, leading to an increase of
management costs of touristic harbours and marinas, and with an impact on the quality of touristic
activities.
Arthropoda:Crustacea
Acartia (Acanthacartia) tonsa Dana, 1849
Acartia (Acanthacartia) tonsa is an opportunistic crustacean species that can dominate
zooplankton communities and displace native copepods (Katsanevakis et al. 2014a; Camatti et al.
2019). On the other hand, A. (Acanthacartia) tonsa has been indicated to intensely graze on algal
blooms (Leppäkoski et al. 2002; Katsanevakis et al. 2014a). A. tonsa has also been used as an
effective indicator species for assessing sediment quality (Buttino et al. 2018).
Callinectes sapidus Rathbun, 1896
The blue crab Callinectes sapidus is native of the western Atlantic and was introduced in European
waters probably through ballast waters; it appeared in the Mediterranean Sea at the mid of the past
century and in the past decades it has spread throughout the basin in coastal and transitional
ecosystems (Mancinelli et al. 2017, 2021). In the eastern Mediterranean and the NW Adriatic (C.
Frolia pers. commu) it is commercially exploited by fishers in coastal waters and lagoons (Çinar
et al. 2005; Bilecenoglu et al. 2013; Abdel Razek et al. 2016; Zotti et al. 2016; Turan et al. 2016;
39
Kevrekidis and Antoniadou 2018; Manfrin et al. 2018; Mehanna et al. 2019). C. sapidus is
considered a pest for fisheries as it damages fishing gear with its claws and preys on trapped catch
(Perdikaris et al. 2016; Mancinelli et al. 2017; Kampouris et al. 2019). Furthermore, the
establishment of the blue crab near areas of mussel and clam aquaculture, poses a threat for the
shellfish production, since C. sapidus preys on these species (Prado et al. 2020b). Abundant
populations of the blue crab may exert significant predation pressure on native biodiversity and
compete with native predators (Gennaio et al. 2006; Carrozzo et al. 2014; Mancinelli et al. 2016;
Pla Ventura et al. 2018; Kampouris et al. 2019). C. sapidus preferentially feeds on mollusks,
crustaceans and fishes, exhibiting a wide spectrum of predation strategies (Rady et al. 2018;
Kampouris et al. 2019); however, it can shift its diet according to resource availability, preying on
animal organisms as well as on living and non-living primary producers simultaneously, and thus
impacting invaded ecosystems at multiple trophic levels (Mancinelli et al. 2013). C. sapidus has
been assessed as a biomonitoring species for heavy metal and microplastic pollution in coastal
marine ecosystems (Genç and Yilmaz 2017; Renzi et al. 2020; Salvat-Leal et al. 2020). Wasted
blue crab shells have the potential to be recycled and reused in innovative applications such as
smart nanocarriers (Nekvapil et al. 2019). The species is under consideration for inclusion in the
List of Invasive Alien Species of Union concern under Regulation (EU) 1143/2014 (Zenetos et al.
2020c).
Dyspanopeus sayi (Smith, 1869)
The xanthid crab Dyspanopeus sayi, native to the western Atlantic, was first recorded from the
Mediterranean Sea in 1992, in the Venice lagoon (Adriatic Sea) (Froglia and Speranza 1993). Ever
since, the species has been found in many transitional ecosystems throughout the Mediterranean
(Ulman et al. 2017). When established and abundant the invasive crab could potentially pose a
threat for biodiversity and the aquaculture industry as an effective predator of mussels and oysters
and a potential competitor of the native crab Carcinus aestuarii Nardo, 1847 for space and
resources (Cabiddu et al. 2020). Dyspanopeus sayi also preys effectively on medium-small sized
(15 – 25 mm shell length) Arcuatula senhousia mussels and prefers the invasive mytilid over the
manilla clam Ruditapes philippinarum (Mistri 2004b).
Erugosquilla massavensis (Kossmann, 1880)
The Massawan mantis shrimp Erugosquilla massavensis is a Lessepsian migrant introduced in the
Mediterranean in 1933 and first recorded as Squilla africana Calman, 1917 (Steuer 1936). The
invasive stomatopod is widespread in the eastern Mediterranean and in Tunisian waters (Ounifi-
Ben Ammor et al. 2015). The species has flourished at some locations and its presence has been
associated with competition and displacement of Squilla mantis (Linnaeus, 1758) (Streftaris and
Zenetos 2006; Özcan et al. 2008), a native stomatopod of commercial importance (Lewinsohn and
Manning 1980). E. massavensis has acquired commercial value and is being locally exploited
(Gianguzza et al. 2019; Koulouri et al. 2020). Furthermore E. massavensis extracts have exhibited
antioxidant activity against toxic metabolites (Fahmy and Hamdi 2011). Its shells have a potential
for the production of bioactive compounds to be used for pharmaceutical and food industry
applications (Abouzeed et al. 2015).
40
Matuta victor (J.C. Fabricius, 1781)
Matuta victor was firstly reported from the Mediterranean in 2012 from Haifa Bay and ever since
has rapidly increased its population in the Levantine Sea (Galil and Mendelson 2013; Crocetta et
al. 2015a; Innocenti et al. 2017). The invasive crab has exhibited intraspecific competitive
behavior for food (Innocenti et al. 2017), yet no impacts on biodiversity or ecosystem services
have been reported from Mediterranean ecosystems.
Metapenaeus monoceros (J.C. Fabricius, 1798)
The speckled shrimp Metapenaeus monoceros is a Lessepsian invader that has established
abundant populations in the eastern Mediterranean and Tunisia (Zenetos et al. 2010; Fiorentino et
al. 2013). M. monoceros has become a major commercial species for Levantine Sea and the
Tunisian plateau fisheries (Ben Abdallah et al. 2003; Can et al. 2004; Mehanna and Usama Khalifa
2007; McCall 2008; Ben Hadj Hamida-Ben Abdallah et al. 2009; Yilmaz et al. 2009; Galil 2011;
Ateş et al. 2013; Otero et al. 2013; Özbilgin et al. 2015; Ulman et al. 2015; Gökçe et al. 2016a;
Turan et al. 2016). The spread of M. monoceros in Tunisian waters raised concerns to local
fisheries due to the potential competitive impact on the commercially important native caramote
prawn Penaeus kerathurus (Forskål, 1775) (Chaouachi et al. 1998; Galil 2007b; Manfrin et al.
2018). Following the introduction of the speckled shrimp in the gulf of Gabes, P. kerathurus
exhibited population declines that were attributed to a competitive interaction for resources (Ben
Abdallah et al. 2003; Hattab et al. 2013). Furthermore, in the Levantine Sea, the species is
associated with negative effects on native shrimps (Saygu et al. 2020). Muscle tissue analysis of
proteins, carbohydrates, lipids, and acids has assessed M. monoceros as beneficial for human
consumption (Ayas et al. 2013; Yerlikaya et al. 2013; Banu et al. 2016). In addition, its edible
parts contain high concentrations of carotenoid content, which gives the speckled shrimp an
exceptional nutritional value as seafood (Yanar et al. 2004). By-products of the speckled shrimp
were investigated and exhibited both antioxidant activities and enzyme inhibitory potential
(Mechri et al. 2020).
Metapenaeus stebbingi Nobili, 1904
The penaeid prawn Metapenaeus stebbingi is a Lessepsian invader, established in the eastern
Mediterranean and Tunisia (Zenetos et al. 2010). M. stebbingi is edible and constitutes an
important biological resource for Levantine fisheries (Wadie and Abdel Razek 1985; Kumlu et al.
1999; Can and Demirci 2004; Can et al. 2004; Mehanna and Usama Khalifa 2007;Otero et al.
2013; Turan et al. 2016). There are uncertainties regarding the impacts of M. stebbingi on the
biodiversity of Mediterranean ecosystems, but it has been speculated that it may compete with the
native prawn Penaeus kerathurus (Otero et al. 2013; Manfrin et al. 2018). At the same time, it
constitutes a preferred prey item for the common stingray Dasyatis pastinaca (Linnaeus, 1758)
(Yeldan et al. 2009). Furthermore, chitosan extracted from M. stebbingi has been tested as an
organic preservative for fish under refrigerated storage (Küçükgülmez et al. 2013; Yanar et al.
2013). The species has also been utilized as an indicator of heavy metal pollution (Saadia et al.
2011).
Paracerceis sculpta (Holmes, 1904)
41
The alien isopod Paracerceis sculpta was firstly reported in the Mediterranean from the bay of
Tunis (Rezig 1978) and ever since has established abundant populations throughout the basin
(Katsanevakis et al. 2014b). Evidence suggest that the species has become a pest in Pialassa
Baiona, a lagoon of the northern Adriatic, with negative competitive effects on native species
(Vincenzi et al. 2013). No impacts have been recorded on ecosystem services in the Mediterranean.
Penaeus aztecus Ives, 1891
The northern brown shrimp Penaeus aztecus, native to western Atlantic, was first reported in the
Mediterranean from Antalya Bay, Turkey, in 2009 (Deval et al. 2010). In the next few years, the
alien shrimp was reported across the Mediterranean basin (Galil et al. 2017). P. aztecus has become
a commercially exploited species in the Levantine and the Ionian Sea (Mytilineou et al. 2016;
Turan et al. 2016; Zava et al. 2018; Kampouris et al. 2018; El-Deeb et al. 2020), although it is
often sold mixed with other penaeid shrimps (Bakir and Aydin 2016; Zava et al. 2018; T.E.
Kampouris, pers. comm.). P. aztecus is a commercially cultured species in Egypt, where postlarvae
are collected during their migration to shallow waters (Sadek et al. 2018). P. aztecus establishment
in Mediterranean ecosystems has raised the awareness of scientists for the potential impact on its
native congeneric Penaeus kerathurus (Kevrekidis 2014).
Penaeus pulchricaudatus Stebbing, 1914
The penaeid shrimp Penaeus pulchricaudatus, previously identified as Penaeus japonicus Spence
Bate, 1888, has become abundant in the eastern Mediterranean basin (Zenetos et al. 2010). In the
past, under the name Penaeus japonicus, two species have been confused: Penaeus japonicus
Spence Bate, 1888, only found in China and Japan Seas, and Penaeus pulchricaudatus Stebbing,
1914, native of western Indian Ocean, southeastern Asia and Australia, that reached the East
Mediterranean via the Suez canal (Tsoi et al. 2007, 2014). These penaeid shrimps are commercially
exploited and highly prized worldwide with an impressive aquaculture farming production (Tsoi
et al. 2014). In the eastern Mediterranean, P. pulchricaudatus has been an important target of
fisheries that significantly contributes to local economy (Mehanna et al. 2005, 2011; Galil 2007b;
Duruer et al. 2008; Edelist et al. 2011; Otero et al. 2013; Corsini-Foka et al. 2015; Ulman et al.
2015; Katsanevakis et al. 2018; Ammar 2019; van Rijn et al. 2020). Pond aquaculture of the
species has also been developed around various locations of the Mediterranean (Galil and Zenetos
2002; Savini et al. 2010; Otero et al. 2013; Manfrin et al. 2018). In the early trials of shrimp
aquaculture in Italy, the released post-larvae were imported from Japan, i.e. P. japonicus (Lumare
1984); in other cases post-larvae were obtained from broodstock imported from the eastern
Mediterranean, i.e. P. pulchricaudatus (Stentiford and Lightner, 2011). The species has a high
nutritional value (Ayas et al. 2013; Yerlikaya et al. 2013). P. pulchricaudatus has been associated
with negative effects on native shrimps (Galil and Zenetos 2002; Otero et al. 2013; Saygu et al.
2020). Furthermore, farm ponds of the invasive prawn are potential points of introduction of
diseases to native fauna (Stentiford and Lightner 2011; Otero et al. 2013; Manfrin et al. 2018).
Penaeus semisulcatus De Haan, 1844 [in De Haan, 1833-1850]
The green tiger shrimp Penaeus semisulcatus is a penaeid species of Indo-Pacific origin that has
been introduced and established in the Mediterranean basin (Zenetos et al. 2010). The introduction
42
of the alien shrimp has been favourable for local economies since the species is fished and
contributes significantly to food provision (Duruer et al. 2008; Galil 2011; Mehanna et al. 2011;
Cataudella et al. 2015; Özbilgin et al. 2015; Ulman et al. 2015; Gökçe et al. 2016b; Turan et al.
2016; El-Gendy et al. 2018; Ammar 2019). At some occasions there have been reports of
overfishing of the species (Manasirli et al. 2014). P. semisulcatus meat is considered a nutritious
food source and optimal for human consumption (Diler and Ataş 2003; Yanar et al. 2004;
Yerlikaya et al. 2013; Banu et al. 2016). Its presence in the Levantine Sea has been associated with
negative effects on native shrimps (Saygu et al. 2020). P. semisulcatus has been farmed in some
Mediterranean locations as it has shown some potential for commercial aquaculture (Browdy and
Samocha 1985; Browdy et al. 1986; Seidman and Issar 1988; Kumlu et al. 2003, 2010, 2019;
Türkmen 2007; Sadek 2010; Hussain et al. 2015). However, shrimp farming is still challenging
for the aquaculture industry and many constraints exist both at production and managerial level
(Colorni 1989; Kumlu et al. 2000; Kir et al. 2004; Sadek 2010;Stentiford and Lightner 2011; Aly
et al. 2021). P. semisulcatus has been utilized as a biomonitoring species for tracing heavy metals
in the marine environment (Kargin et al. 2001; Çoğun et al. 2005; Yilmaz and Yilmaz 2007; Firat
et al. 2008; Çiftçi et al. 2021).
Percnon gibbesi (H. Milne Edwards, 1853)
The grapsoid crab Percnon gibbesi was firstly reported from the Mediterranean in 1999. Ever
since, abundant populations have been recorded throughout the Mediterranean, where favourable
conditions are met (Garcia and Reviriego 2000; Relini et al. 2000; Sciberras and Schembri 2008;
Raineri and Savini 2010; Katsanevakis et al. 2011; Stasolla et al. 2016). Nevertheless, the vector
of introduction within the basin remains uknown and no strong evidence have revealed an alien
origin of the species (Mannino et al. 2017b). On the contrary, there have been speculations that P.
gibbesi may have spread within the Mediterranean by a natural expansion (Abelló et al. 2003). As
such, P. gibbesi should be considered as cryptogenic (Mannino et al. 2017b). P. gibbesi is a
herbivorous species that inhabits shallow rocky bottoms and shares a diet overlap with a variety
of native invertebrates such as the native grapsid Pachygrapsus marmoratus (J.C. Fabricius,
1787), the xanthoid Eriphia verrucosa (Forsskål, 1775) and the echinoids Paracentrotus lividus
and Arbacia lixula (Linnaeus, 1758) (Sciberras and Schembri 2008; Katsanevakis et al. 2014a;
Félix-Hackradt et al. 2018). However, no significant negative impacts against native biodiversity
have been reported to be caused by this P. gibbesi, neither by competition nor grazing (Sciberras
and Schembri 2008; Katsanevakis et al. 2014a; Guillén et al. 2016; Félix-Hackradt et al. 2018).
The fish Gobius paganellus Linnaeus, 1758 feeds on small benthic crustaceans and P. gibbesi has
become a preferred prey for this native rocky goby (Tiralongo et al. 2021). When abundant and
diverse, native predators have shown a potentially effective capability in controlling the
populations of P. gibbesi with intensive predation pressure (Noè et al. 2018b).
Portunus segnis (Forskål, 1775)
The blue swimming crab Portunus segnis, previously identified in the Mediterranean Sea as
Portunus pelagicus (Linnaeus, 1758), is an early Lessepsian invader of Indian Ocean origin that
is well-established in the eastern and southern-central part of the basin and has also been recorded
in the northern Tyrrhenian Sea (Fox, 1924; Crocetta 2006; Zenetos et al. 2010; Annabi et al. 2018)
43
and northern Tunisia (Mili et al. 2020; Shaiek et al. 2021). P. segnis has become an exploitable
biological resource for local fisheries (Fox 1924; Galil 2000; Corsini-Foka et al. 2015; Gökçe et
al. 2016a; Turan et al. 2016; Galil et al. 2018; Katsanevakis et al. 2018; Ben Abdallah-Ben Hadj
Hamida et al. 2019a; van Rijn et al. 2020) with an adequate nutritional value for human
consumption (Bejaoui et al. 2017; Olgunoğlu and Olgunoğlu 2017; Artar and Olgunoğlu 2018).
Nevertheless, the invasive crab preys on captured commercial species in fishing gears and is
entangled in nets in great numbers, causing economic losses to small scale fisheries (Crocetta et
al. 2015a; Galil et al. 2018; Khamassi et al. 2019). P. segnis is an opportunistic carnivore (Annabi
et al. 2018), preying mainly on crustaceans, mollusks and fish (Ben Abdallah-Ben Hadj Hamida
et al. 2019b). P. segnis viscera have been assessed as a potential alternative resource of valuable
bio-products for the food industry (Hamdi et al. 2017; Maalej et al. 2021).
Rhithropanopeus harrisii (Gould, 1841)
Rhithropanopeus harrisii has established populations in locations of the Adriatic Sea, the central
and the western Mediterranean (Zenetos et al. 2010; Langeneck et al. 2015; Ferrario et al. 2017);
nevertheless, no documented impact on biodiversity and ecosystem services has been recorded for
this species to date in Mediterranean ecosystems.
Echinodermata
Diadema setosum (Leske, 1778)
The long-spined urchin Diadema setosum is an alien species that was introduced in the
Mediterranean in 2006 in the southern coasts of Turkey (Yokeş and Galil 2006). It hides in cryptic
habitats such as rock crevices during the day and forages at nearby rocky beds at night (Yokeş and
Galil 2006). Currently, this alien echinoid is established in the Levantine Sea, the Aegean and the
Ionian Sea (Ragkousis et al. 2020). In the Dodecanese islands of the Aegean Sea, it has formed
sparse populations with some local dense patches (Vafidis et al. 2021). Reduction of native
echinoids in the eastern basin (Yeruham et al. 2015; Yeruham et al. 2020) may potentially enhance
the invasion of D. setosum through an increase in empty niches availability (Bronstein et al. 2017).
No impacts of the alien echinoid were found on Mediterranean native biodiversity and ecosystem
services. Nevertheless D. setosum poses a threat for human health (Yokeş and Galil 2006; Galil
2018; Bédry et al. 2021). Its spines contain venom and are able to pierce swimmers, snorkelers
and divers, releasing the venom inside the human flesh and causing a variety of symptoms that
may last from a few hours to days (Galil 2018; Bédry et al. 2021). Spines are difficult to remove
from inside the body and healing may require weeks to complete (Yokeş and Galil 2006).
Occasionally more serious cases have been reported (Galil 2018; Bédry et al. 2021).
Synaptula reciprocans (Forsskål, 1775)
With an Indo-pacific origin, Synaptula reciprocans is a holothurian that has been introduced in the
eastern Mediterranean for almost five decades (Cherbonnier 1986; Galil 2007a; Zenetos et al.
2010). During the more recent years the species has expanded its range in the eastern basin and
44
has established abundant populations in shallow rocky and sandy bottoms (Antoniadou and Vafidis
2009; Ragkousis et al. 2017). No impact has been reported as of yet of S. reciprocans on
biodiversity or ecosystem services in Mediterranean ecoregions.
Chordata:Ascidiacea
Botrylloides violaceus Oka, 1927
The colonial ascidian Botrylloides violaceus is considered a major nuisance fouling species
worldwide, with impacts in aquaculture facilities and natural ecosystems in North America (Bock
et al. 2011 and references therein; Katsanevakis et al. 2014a). No impacts of this alien colonial
ascidian have been reported as of yet in Mediterranean ecosystems.
Botrylloides diegensis Ritter & Forsyth, 1917
No impacts of this alien colonial ascidian have been reported as of yet in Mediterranean
ecosystems.
Ciona robusta Hoshino & Tokioka, 1967
Ciona robusta is considered an introduced solitary ascidian in the Mediterranean established for
more than a century (Zenetos et al. 2017), originally described from Japan. Up until 2015, records
of this introduced tunicate in the Mediterranean were classified as Ciona intestinalis (Linnaeus,
1767) (Brunetti et al. 2015; Pennati et al. 2015). C. robusta is a suspension filter-feeder ascidian
with a great potential to be used as a biomonitoring species of heavy metal pollution (Arienzo et
al. 2014; Laura et al. 2021). It has been observed to foul aquaculture facilities in dense aggregations
(Chebbi et al. 2010). C. robusta has also been used to clarify the molecular mechanisms of its
innate immune system response to bacterial infection (Arizza et al. 2020). Ciona species may also
be a source of biofuel, cellulose and other bioresources (Zhao and Li 2014, 2016; Hruzova et al.
2020) but there is no such use yet in the Mediterranean. No impacts by C. robusta on
Mediterranean biodiversity have been documented so far.
Clavelina oblonga Herdman, 1880
Clavelina oblonga is a species native to the southern Atlantic coast of North America and the
Caribbean, and has been introduced in South America, in the Atlantic archipelagos of Azores and
Cape Verde, and in Senegal (Rocha et al. 2012). In the Mediterranean, this species has been
previously known as Clavelina phlegraea Salfi, 1929 and found in Italy (Mastrototaro and Tursi
2010) and Corsica (Monniot et al. 1986), always in confined environments. Its taxonomic status
was clarified by Ordóñez et al. (2016), who reported an outbreak of this species smothering bivalve
cultures in the Iberian littoral. This ascidiacan can be very abundant in aquaculture settings, fouling
shellfish and equipment (Mastrototaro et al. 2008; Ordóñez et al. 2016; Casso et al. 2018). A new
antifungal agent has been found in extracts of Clavelina oblonga (Kossuga et al. 2004).
Didemnum vexillum Kott, 2002
45
The invasive ascidian Didemnum vexillum is a colonial didemnid native to the Northwestern
Pacific ocean. D. vexillum is globally considered highly invasive due to its ability to colonize
extended substrate surfaces, overgrowing native organisms and causing severe impacts to benthic
communities (Katsanevakis et al. 2014a; Mckenzie et al. 2017). In the western Mediterranean Sea
it has been recorded overgrowing oysters and their associated epifauna in aquaculture facilities,
causing economic losses (Ordóñez et al., 2015; Casso et al., 2019). It has been pointed out that this
species can act as a reservoir of molluscan pathogens in these environments (Costello et al. 2020).
Herdmania momus Savigny, 1816
The Lessepsian migrant Herdmania momus is a common solitary ascidian of the Red Sea
established in the eastern Mediterranean basin (Zenetos et al. 2010). During the last two decades,
this alien ascidian has increased its population numbers in the Levantine Sea, successfully
colonizing both artificial and natural substrates (Gewing et al. 2014, 2017). Temperature
constitutes an important environmental factor for the distribution of the species and the ongoing
seawater warming within the Mediterranean could facilitate its expansion (Gewing et al. 2019).
Due to its nature as a filter-feeder organism, H. momus has been used as a biological pollution
indicator of phthalate acid esters, microplastic particles and pharmaceutically active compounds
in the water column (Vered et al. 2019; Navon et al. 2020).
Microcosmus squamiger Michaelsen, 1927
The alien solitary sea squirt Microcosmus squamiger is established in the western and central
Mediterranean basin (Zenetos et al. 2010). M. squamiger can become invasive and dominant on
available artificial or natural surfaces in shallow hard substrates through the formation of
monospecific dense aggregations, outcompeting native species and completely altering the
invaded habitat (Turon et al. 2007; Mastrototaro and Dappiano 2008; Rius et al. 2009; Ordóñez et
al. 2013; Otero et al. 2013; Garcia et al. 2015). This invasive ascidian is also capable of fouling
mussel cultures, inhibiting their production (Otero et al. 2013). The native edible gastropod
Stramonita haemastoma has been observed preying on M. squamiger and has exhibited a positive
correlation with the invasive ascidian (Rius et al. 2009).
Polyandrocarpa zorritensis (Van Name, 1931)
This species was described from Peru and is now globally distributed (Monniot 2018). Since the
1970s it has been found in the Mediterranean, where it is established in the Western and Central
areas (Brunetti and Mastrototaro 2004; Zenetos et al. 2010). It is abundant in enclosed
environments and aquaculture settings (Mastrototaro et al. 2008), potentially interfering with the
farming activities. It can also be found in natural habitats (Mastrototaro et al. 2008; García et al.
2015). P. zorritensis possesses species-specific microbiota that can contribute to its expansion
(Evans et al. 2017). According to Stabili et al. (2015, 2016b), it can also contribute to mitigating
microbial pollution with its filter-feeding activity.
Styela plicata (Lesueur, 1823)
Styela plicata is a solitary ascidian whose origin, although not clearly defined as yet (Pineda et al.
2011), is considered to be the NW Pacific (Barros et al. 2009). It is widely distributed in tropical
46
and temperate waters and is present throughout the Mediterranean Sea (Locke 2009). This invasive
tunicate is capable of colonizing in dense aggregations equipment used in aquaculture farms and
also fouls farmed mussels, inhibiting production and causing economic losses (Chebbi et al. 2010;
Antoniadou et al. 2013; Casso et al. 2018; Petović et al. 2019; Pica et al. 2019). S. plicata has the
ability to dominate the available substrate at eutrophic environments and even outcompete the
native Mytilus galloprovincialis (Pica et al. 2019). On the other hand, dense fouling aggregations
of S. plicata filter-feed on water column particles and contribute to the removal of suspended
organic material at locations of sub-optimal environmental conditions (Montalto et al. 2020). The
invasive tunicate has also been assessed as a prominent bio-monitoring indicator of heavy metal
pollution and pharmaceutical contamination in coastal waters (Pineda et al. 2012; Aydın-Önen
2016; Bellante et al. 2016; Navon et al. 2020). Crude extracts from S. plicata specimens from the
coastal lake Faro in Italy exhibited a significant antimicrobial activity in laboratory experiments
(Palanisamy et al. 2016). Aqueous extracts of the species, obtained from the port of Alexandria,
have shown potential for chemotherapeutic drugs production against colon cancer (Salim et al.
2020). This species has been proved to remove the bacterium Vibrio alginolyticus (Miyamoto et
al. 1961) Sakazaki, 1968, purifying contaminated water through its filter-feeding capability
(Stabili et al. 2016b). Furthermore, it can also be a source of cellulose and other bioresources (Zhao
and Li 2014, 2016). The closely related species Styela clava Herdman, 1881 has been shown to
possess interesting anti-inflammatory and anti-aging properties of potential use in new cosmetics
(Lee et al. 2015).
Chordata:Osteichthyes
Alepes djedaba (Forsskål, 1775)
The shrimp scad, Alepes djedaba is a medium size pelagic carangid with an Indo-Pacific
distribution, that was first recorded in the Mediterranean by Steinitz (1927) as Caranx calla. Since
then, A. djedaba has extended its distribution and spread to the eastern and central Mediterranean
basin and even reached the Black Sea (Turan et al. 2017). Juveniles of A. djedaba are frequently
found sheltering under the umbrellas of native (Artüz and Tunçer 2017) and invasive
scyphomedusae (Galil et al. 1990). Although it is not a targeted carangid species, it is common in
the Levantine fisheries, being caught in large quantities and with various fishing methods (Goren
and Galil 2005; Golani et al. 2013a; Otero 2013; Ali 2018; Ragheb et al. 2019; Shakman et al.
2019).
Apogonichthyoides pharaonis (Bellotti, 1874)
Introduced in the eastern Mediterranean in 1946, the Pharaoh cardinal fish is a nocturnal fish whose
presence and abundance might be facilitated by the lack of competitors (Haas and Steinitz 1947;
Otero et al. 2013). No impacts on biodiversity or ecosystem services have been documented so far
for this species.
Atherinomorus forskalii (Rüppell, 1838)
47
Atherinomorus forskalii is the first Lessepsian fish recorded in the Mediterranean Sea (Tillier
1902; Irmak and Özden 2020). Today the species has managed to establish abundant populations
throughout the eastern and southern sectors of the basin. A. forskalii is a coastal pelagic feeder,
characterized by a wide trophic spectrum, ranging from pelagic gastropods, copepods to juvenile
fish and insects, and potentially competing with other small pelagic predators such as the natives
Atherina boyeri Risso, 1810 and Belone belone (Linnaeus, 1760) (Irmak and Özden 2020).
Nevertheless, ecological impacts have not been investigated so far. Rigid scales, hiding in
sheltered areas, and other behavioral and morphological traits may provide a better anti-predatory
defense to A. forskalii with respect to the native A. boyeri (Irmak and Özden 2020). Still, when
present at large schools, the species could become an important prey for coastal predators (Otero
et al. 2013). Due to its small size A. forskalii is not a preferred fish catch of the fisheries industry,
but it is currently marketed in Egypt, where it is particularly abundant (Ali 2018; Otero et al. 2013).
Decapterus russelli (Rüppell, 1830)
The Indian scad, Decapterus russelli is yet another medium sized pelagic species that invaded the
Mediterranean Sea where it presently has established populations. First observed at the end of
2005 in Israel (Golani 2006), it is currently known from as far west as the Mersin bay in Turkish
Mediterranean coast (Sakinan and Örek 2011). Nevertheless, the close resemblance to other native
carangids, such as Trachurus or Caranx spp. may cause overlooking of its true distribution and
abundance in the Mediterranean (Golani 2006). D. russelli is known to prey mainly on fish in the
Levant, whereas crustaceans are more important in its diet in the Arabian Sea (Gilaad et al. 2017).
Currently, it constitutes an important component of the Israeli bottom trawl and coastal purse seine
fisheries (Katsanevakis et al. 2018). It is marketed in combination with the catch of horse-
mackerels Trachurus spp.
Dussumieria elopsoides Bleeker, 1849
Being one of the first invasive clupeids, the slender rainbow sardine Dussumieria elopsoides has
been known in the Levant Basin even before the middle of the 20th century (Ben-Tuvia 1953).
Nevertheless, it currently provides catches of limited economic importance (Galil 2007a; Turan
2010; Ali 2018) that are occasionally mixed with other small pelagic species (Çinar et al. 2005).
Etrumeus golanii Di Battista, Randall & Bowen, 2012
Etrumeus golanii is a pelagic fish of Lessepsian origin, firstly recorded in the Mediterranean in
1961, from the coasts of Israel. Currently, the species has been reported from many Mediterranean
countries such as Egypt, Cyprus, Turkey, Greece, Tunisia, Italy, Algeria and Morocco, up to the
western end of the Mediterranean (Galil et al. 2018). The Mediterranean population of Etrumeus
golanii was historically misidentified with Etrumeus teres until the description of E. golanii as a
new species (Di Battista et al. 2012). The species is being exploited in eastern Mediterranean
countries, where it has become a common catch (Corsini-Foka et al. 2014; Shakman et al. 2019;
Çinar et al. 2021).
Fistularia commersonii Rüppell, 1838
48
The bluespotted cornetfish Fistularia commersonii, of Lessepsian origin, is one of the most
successful invaders of the Mediterranean. The first Mediterranean record of F. commersonii was
reported from Israel in early 2000 (Golani 2000). Today the species has expanded its distribution
throughout the entire Mediterranean, with impressive spread rates (Karachle et al. 2004; Golani et
al. 2007; Azzurro et al. 2013). The bluespotted cornetfish is a voracious carnivore that consumes
a wide variety of native fish, including species of commercial value such as Spicara smaris
(Linnaeus, 1758), Boops boops (Linnaeus, 1758) and Mullus spp. (Kalogirou et al. 2007; Bariche
et al. 2009; Castriota et al. 2012; Otero et al. 2013; Saad and Khalaf 2016; Karachle and Stergiou
2017; Mouine-Oueslati et al. 2017). One of the factors that have led to F. commersonii invasion
success could be attributed to prey naivety of Mediterranean fish against this novel predator (Sih
et al. 2010; D’Amen and Azzurro 2020). Current knowledge and projections indicate the potential
for F. commersonii to impact the community structure of the invaded ecosystem (Kalogirou et al.
2007; Otero et al. 2013; Pinnegar et al. 2014) but specific studies are currently missing. At the
same time, some seabirds and large pelagic fish are supposed to prey on F. commersonii, which
would represent a novel food resource (Pinnegar et al. 2014). Modelling the trophic web of Cyprus
by Ecopath with Ecosim, Michailidis (2021) found that an increase of the bluespotted cornetfish
biomass would hugely benefit large pelagic fish (by as much as 280%) especially in scenarios of
reduced overall fishing pressure. F. commersonii also includes in its diet invasive fish such as
small Siganus rivulatus Forsskål & Niebuhr, 1775 and Pterois miles (Bennett, 1828), indicating
thus a potential to control the population of other invaders (Galanidi et al. 2018; Giakoumi et al.
2019; Michailidis et al. 2019; Michailidis 2021). Furthermore, as an effective novel predator, F.
commersonii has the potential to compete with native predators that hunt the same prey (Otero et
al. 2013; Fanelli et al. 2015). At most countries’ fisheries, the invasive bluespotted cornetfish
constitutes a by-catch fish, but at some localities of the eastern Mediterranean F. commersonii has
acquired economic value due to its white palatable free of spines flesh (Otero et al. 2013; Corsini-
Foka et al. 2017; Ali 2018; Shakman et al. 2019).
Herklotsichthys punctatus (Rüppell, 1837)
Another veteran Lessepsian clupeid that was reported from the Mediterranean side of the Suez
Canal before 1943 from an unknown location by Bertin (1943). The spotback herring,
Herklotsichthys punctatus, is currently known throughout the eastern Levant, having little to
moderate importance in the catch of coastal pelagic fisheries (Galil 2007a; Turan 2010; Ali 2018).
Lagocephalus sceleratus (Gmelin, 1789)
The silver-cheeked toad-fish Lagocephalus sceleratus is a Lessepsian invader, first recorded in the
Mediterranean from Turkish coasts in 2003 (Filiz and Er 2004; Akyol et al. 2005). In less than two
decades L. sceleratus has managed to form abundant populations in the eastern Mediterranean,
expanding its distribution up to the western basin, at the doorstep of the Atlantic Ocean (Azzurro
et al. 2020). Distribution models indicate that the species is being favoured by suitable climatic
conditions (Coro et al. 2018; D’Amen and Azzurro 2019). L. sceleratus is a predator that feeds on
invertebrates and fish (Kalogirou 2013; Giakoumi et al. 2019). Large individuals of L. sceleratus
prey on the commercially important cephalopods Sepia officinalis Linnaeus, 1758 and Octopus
49
vulgaris Cuvier, 1797 (Pancucci-Papadopoulou et al. 2012; Kalogirou 2013; Akbora et al. 2020;
Hussain et al. 2020). L. sceleratus feeds indistinctively on native and alien fish and could
potentially exert population control of other invasive fish through its predation (Giakoumi et al.
2019). Nevertheless, when modelling the coastal food-web of Cyprus, based on an Ecopath with
Ecosim model, Michailidis (2021) found an increase of siganids when pufferfish biomass was
forced to increase in the model. Using a similar modelling approach, Saygu et al. 2020 depicted
negative effects of L. sceleratus on native demershal fish with high retention rate. L. sceleratus
has a significant negative impact on fisheries in many coastal areas, damaging fishing gear and
catch, and fishers had to adjust their fishing practices (gear, depths, time of the day, etc.) in order
to avoid the species (Katsanevakis et al. 2009; Rousou et al. 2014; Boustany et al. 2015; Ünal et
al. 2015; Ünal and Bodur 2017; Abd Rabou 2019; Akbora et al. 2020). The invasion of L.
sceleratus constitutes a major threat to Mediterranean fisheries with major economic losses (Coro
et al. 2018). Furthermore, L. sceleratus contains a strong paralytic natural neurotoxin known as
tetrodotoxin (TTX) in its skin, tissues and internal organs, which can be lethal for humans or cause
a variety of severe symptoms (Noguchi and Ebesu 2001; Bentur et al. 2008; Eisenman et al. 2008;
Katikou et al. 2009; Awada et al. 2010; Azzurro et al. 2016; Kosker et al. 2016; Acar et al. 2017;
Katikou and Vlamis 2017; Rambla-Alegre et al. 2017; Galil 2018; Abd Rabou 2019; Kosker et al.
2019; Leonardo et al. 2019; Akbora et al. 2020; Ujević et al. 2020; Uysal and Turan 2020; Bédry
et al. 2021). There have also been incidents of L. sceleratus attacks on swimmers with seriously
inflicted wounds and amputations, caused by thier sharp bite (Sümen and Bilecenoğlu 2019). Its
marketing and consumption is currently banned in the EU (Regulation 854/2004/EC) and most
Mediterranean countries, though some illegal marketing exists in some countries such as Egypt,
Syria, Lebanon and Turkey.
Nemipterus randalli Russell, 1986
Randall’s threadfin bream Nemipterus randalli is a recent Lessepsian invader, well-established in
the eastern basin (Golani and Sonin 2006). It is considered a valuable biological resource for many
regional fisheries of the Levantine Sea, where the species is abundant and constitutes an important
portion of the catches (Eryaşar et al. 2014; Stern et al. 2014; Yemisken et al. 2014; Iglésias and
Frotté 2015; Katsanevakis et al. 2018; Demirci et al. 2018; Dalyan 2020; van Rijn et al. 2020;
Çinar et al. 2021). Abundant N. randalli populations can induce predation pressure to populations
of benthic decapod crustaceans (Otero et al. 2013). Moreover, N. randalli can compete with native
species for food resources (Yapici and Filiz 2019) and is suspected to have displaced the native
Pagellus erythrinus (Linnaeus, 1758) from the shallow waters of Israel (Edelist et al. 2013). N.
randalli has also been effectively used as a bioindicator of microplastics and metal concentration
in coastal waters (Güven et al. 2017; Çiftçi et al. 2021).
Parexocoetus mento (Valenciennes, 1847)
An old Lessepsian fish (Bruun 1935) of Indo-Pacific origin, Parexocoetus mento is established in
the eastern and central Mediterranean and the Adriatic Sea (Zenetos et al. 2010). Even though the
alien fish is caught by purse seines, it is of no commercial value (Otero et al. 2013). No impacts
on biodiversity or ecosystem services were traced for this species.
50
Parupeneus forsskali (Fourmanoir & Guézé, 1976)
The goatfish Parupeneus forsskali is a recent Lessepsian invader that has rapidly colonized the
eastern basin. Five years after its record in Cyprus, this invasive species has possibly become the
most abundant mullid in Cypriot shallow waters (Evagelopoulos et al. 2020; K. Tsirintanis, pers.
obs.). It has become the prevalent catch of Cypriot artisanal fisheries among mullids, possibly
indicating a replacement of the native commercial Mullus spp. due to competition (Evagelopoulos
et al. 2020). This is probably attributed to the trophic niche similarities between P. forsskali and
Mullus spp., both consuming small benthic invertebrates (Mahmoud et al. 2017; Evagelopoulos et
al. 2020). The species has a high market value in Cyprus (Evagelopoulos et al. 2020) and it is
currently sold at an average price of 15 euros/Kg (Azzurro pers. obs.).
Pempheris rhomboidea Kossmann & Räuber, 1877
The sweeper Pempheris rhomboidea, of Lessepsian origin, has established abundant populations
in the eastern Mediterranean (Zenetos et al. 2010). For many years, the Mediterranean population
of P. rhomboidea had been misidentified as P. vanicolensis, but a genetic study revealed the true
identity of this Lessepsian invader (Azzurro et al. 2015). P. rhomboidea forms large schools that
inhabit caves and small crevices and may potentially compete for space and resources with native
species that occupy the same habitats, such as the native Apogon imperbis (Linnaeus, 1758) (Goren
and Galil 2005; Bussotti et al. 2015; Michailidis et al. 2019). At daytime these fish shelter inside
caves and crevices, while at night-time they aggregate in groups outside of their shelters to forage
on planktonic crustaceans (Golani and Diamant 1991). Such movements by P. rhomboidea
schools, in and out of their oligotrophic shelters, may affect marine cave ecosystems and their
fragile communities with a flow of organic matter from the outer environment towards the
oligotrophic cave interior (Otero et al. 2013; Gerovasileiou et al. 2016).
Plotosus lineatus (Thunberg, 1787)
The striped eel catfish Plotosus lineatus is a Lessepsian invader that was first detected in the
Mediterranean in 2001 in Israel, and rapidly increased its population during the following years
(Golani 2002; Edelist et al. 2012). Currently, it can be found in the easternmost part of the Levant,
and as north to Iskenderun bay in Turkey (Doğdu et al 2016). P. lineatus is a venomous stinging
fish and represents a threat for human health, in particular harming fishers who try to remove it
with bare hands from nets, and swimmers that accidentally come in contact with it (Gweta et al.
2008; Golani 2010; Bentur et al. 2018; Peyton et al. 2020; Turan et al. 2020; Bédry et al. 2021).
The majority of injuries cause intense pain, swelling, cyanosis, numbness, swelling, erythema,
fever, muscle fasciculations and severe lymphadenopathy (Galil 2018; Bédry et al. 2021). Worse
symptoms like gangrene are rare and probably related to large quantities of injected venom (Galil
2018). P. lineatus causes economic losses in regional fisheries as it constitutes a major discard
species that is difficult to remove from fishing gear (Edelist et al. 2012; Galanidi et al. 2018). The
invasion of the striped eel catfish has been associated with competitive interactions for food
resources and space with native fish such as Mullus spp. and Trachinus spp. (Edelist et al. 2012;
Arndt et al. 2018).
Pterois miles (Bennett, 1828)
51
The lionfish Pterois miles is a successful invader of Lessepsian origin that has caused detrimental
effects on the native ecosystems within its invaded range in western Atlantic (Albins and Hixon
2012; Hixon et al. 2016; Ingeman 2016). In the Mediterranean basin, P. miles was firstly recorded
in Israel in 1991 (Golani and Sonin 1992) but no further records were reported until 2012, when a
rapid population increase was documented in the eastern Mediterranean (Azzurro and Bariche
2017; Dimitriadis et al. 2020). Lionfish are voracious opportunistic ambush predators and their
establishment in Mediterranean ecosystems has the potential to significantly disturb local food
webs (Galanidi et al. 2018; D’Agostino et al. 2020). Many of its prey shows naivety against this
novel predator (Sih et al. 2010; D’Agostino et al. 2020; D’Amen and Azzurro 2020). The invasive
lionfish feed on a variety of fish species, some of which are of commercial value such as Spicara
smaris and some with an important ecological role such as Chromis chromis (Linnaeus, 1758),
and secondarily on other invertebrates (Pinnegar 2018; Zannaki et al. 2019; D’Agostino et al.
2020; Savva et al. 2020). Their prey repertoire also overlaps with that of native predators making
lionfish potential competitors of native consumers (Savva et al. 2020). Lionfish spines are
venomous for humans and stinging by the fish could cause intense pain and swelling that usually
lasts for hours (Galil 2018; Sümen et al. 2018; Uysal and Turan 2020;Bédry et al. 2021).
Inflammatory edema, cyanosis, muscle cramps, anesthesia, paresthesia, tachycardia, bradycardia,
hypertension, hypotension, gastrointestinal disorders, fainting and dizziness, hyperthermia, and
polypnea have also occurred as symptoms (Bédry et al. 2021 and references therein). On the other
hand, P. miles is an impressive and beautiful fish that is appreciated by some people adding a
positive impact on aesthetic values and recreational diving (Jimenez et al. 2017).
Sargocentron rubrum (Forsskål, 1775)
The Indo-Pacific red squirrelfish Sargocentron rubrum is one of the earliest Lessepsian fish of the
Mediterranean first reported from Israel in 1945 and now widespread in the eastern basin (Hass
and Steinitz 1947; Golani and Ben-Tuvia 1985; Zenetos et al. 2010). This invasive species has
become a common catch at many localities (Carpentieri et al. 2009; Ali 2018) and no matter the
value that it has acquired in the fish market, it is still considered underexploited (Farrag et al.
2018). The red squirrelfish hides inside caves and crevices during the day and moves out of their
shelter during the night to feed preferentially on crustaceans, polychaetes, mollusks and fish
(Golani et al. 2013b; Gerovasileiou et al. 2016). Similarly to what has been already discussed for
Pempheris rhomboidea, such movements of S. rubrum, in and out of their oligotrophic cave
shelter, create a flow of organic matter from the outer environment towards the cave interior
potentially affecting the marine cave ecosystem and its biota (Otero et al. 2013; Gerovasileiou et
al. 2016). Other native species such as Epinephelus marginatus and Diplodus sargus could
compete with S. rubrum for space (Spanier 2000).
Saurida lessepsianus Russell, Golani & Tikochinski, 2015
The alien lizardfish Saurida lessepsianus, of Lessepsian origin, was recently recognized as a new
species, previously misidentified as Saurida undosquamis (Richardson, 1848) (Russell et al.
2015). The species was introduced in the Mediterranean during the 1950s and has acquired
commercial value (Galil 2007a and references therein; Keskin et al. 2011; Otero et al. 2013; Çiçek
et al. 2014; Yemisken et al. 2014; Ali 2018; Dalyan 2020). S. lessepsianus is a piscivorous predator
52
that causes negative impacts on native fish through predation (Saygu et al. 2020). It has also been
associated to negative competitive effects on native fish such as the lizard fish Synodus saurus
(Linnaeus, 1758), eels and morays or fish of more significant commercial value such as the native
hake, Merluccius merluccius (Linnaeus, 1758) (Ben Yami and Glaser 1974; Otero et al. 2013;
Katsanevakis et al. 2014a; Gilaad et al. 2017; Michailidis et al. 2019). The invasive fish has also
been used as a biomonitoring indicator for the presence of metals in coastal waters (Çiftçi et al.
2021).
Scomberomorus commerson (Lacepéde, 1800)
The narrow-barred Spanish Mackerel is a schooling migratory pelagic fish that was first reported
from the eastern Mediterranean by Hornell (1935) and can now be seen as westward as Tunisia
(Ben Souissi et al 2006). S. commerson is regarded as a commercial species with high demand and
is considered as one of the most exploited species in the Mediterranean Levantine fisheries (Goren
and Galil 2005; Ali 2018; Katsanevakis et al. 2018; Shakman et al. 2019; Mohsen et al 2020; van
Rijn et al. 2020).
Siganus luridus (Rüppell, 1829); Siganus rivulatus Forsskål & Niebuhr, 1775
The two alien siganids, the dusky spinefoot Siganus luridus and the marbled spinefoot S. rivulatus
are old Lessepsian invaders of the Mediterranean (Steinitz 1927; Gruvel 1931). They dominate
fish communities in many coastal areas, especially in the eastern basin (Bariche et al. 2004;
Katsanevakis et al. 2009, 2011; Thessalou-Legaki et al. 2012; Giakoumi 2014; Sini et al. 2019).
Siganus spp. are herbivorous fish and their introduction in the Mediterranean has caused
detrimental impacts on the Mediterranean macroalgal communities (Bianchi et al. 2014; Giakoumi
2014; Katsanevakis et al. 2014a; Smith et al. 2014; Vergés et al. 2014; Corrales et al. 2017;
Galanidi et al. 2018; Rilov et al. 2018; Yeruham et al. 2020). They overgraze and deplete algal
vegetation to the point that they shift the ecosystem state from algal forests to isolated barren
grounds or turf-dominated communities of low complexity (Sala et al. 2011; Yeruham et al. 2020).
This is a devastating impact to biodiversity since lush algal forests, such as those formed by brown
perennial seaweeds of Cystoseira sensu lato, significantly support marine life as nurseries for a
variety of species, provide habitat for benthic communities, and they are also considered a
threatened biotope in many Mediterranean regions (Cheminée et al. 2013; Otero et al. 2013;
Katsanevakis et al. 2014a). Furthermore, they provide important ecosystem services such as food
provision, biotic materials, climate regulation, water purification, cognitive benefits, recreation,
symbolic and aesthetic values, and life cycle maintenance (Salomidi et al. 2012; Katsanevakis et
al. 2014a). The benefits that these cornerstone communities provide to biodiversity and humans
are heavily threatened by the invasion of the two siganid species in the Mediterranean. The success
of Siganus spp. in the Mediterranean has also occurred at the expenses of the native herbivorous
fish Sarpa salpa that is submitted to the combined effects of climate change and competition with
tropical herbivores (Bariche et al. 2004; Kalogirou et al. 2012a; Otero et al. 2013; Fanelli et al.
2015). Siganus spp. deplete fleshy macroalgae, the favourable food resource of the echinoid
Paracentrotus lividus, contributing additively with ocean warming to the decline of the sea urchin
in the eastern Mediterranean (Yeruham et al. 2020). On the other hand, in the Levantine Sea the
siganids have contributed to the acceleration of algal recycling and constitute an optimal food
53
source for larger predators such as groupers (Goren and Galil 2005). Siganus spp. have acquired a
commercial value at local fish markets at many countries of the eastern Mediterranean (Carpentieri
et al. 2009; Katsanevakis et al. 2009, 2014; Kalogirou et al. 2012a; Shakman et al. 2019; Soykan
et al. 2020), reaching the value of 25 euros/Kg in Cyprus (Azzurro pers. obs). Rabbitfish spines
contain venom that is being released after penetration and causes intense pain to humans (Galil
2018; Uysal and Turan 2020; Bédry et al. 2021). Cases of poisoning and intoxication after the
consumption of siganids have also been reported (Herzberg 1973; Raikhlin-Eisenkraft and Bentur
2002; Bédry et al. 2021).
Sphyraena chrysotaenia Klunzinger, 1884
An alien invasive species of barracuda, first described in Israel by Spicer (1931). An inshore
piscivorous pelagic species that can be found in great abundance in bottom trawl catches (Wadie
and Rizkalla 2001; Katsanevakis et al. 2018). In addition to its economic importance (Rim et al
2007; Kalogirou et al. 2012b; Ali 2018; Shakman et al. 2019; Çinar et al. 2021), this fish has shown
the potential to influence the dynamics of invasive and native parasite-host interactions
(Boussellaa et al. 2018).
Torquigener flavimaculosus Hardy & Randall, 1983
The Yellow spotted Puffer Torquigener flavimaculosus is a Lessepsian invader, firstly recorded in
the Mediterranean from Haifa Bay, Israel (Golani 1987). The species contains tetrodotoxin, a lethal
toxin for humans, in several of its internal organs but also on its skin (Kosker et al. 2018) and thus
its marketing and consumption is prohibited in EU (Regulation 854/2004/EC) and most
Mediterranean countries. Management measures should be considered by local authorities in order
to promote the avoidance of this highly toxic invader by all users of the sea (Kosker et al. 2018).
T. flavimaculosus is a voracious predator that preys indistinctively on native and alien fish
(Giakoumi et al. 2019). The species has also been associated with negative effects on native
demershal fishes with high retention rate (Saygu et al. 2020).
Upeneus moluccensis (Bleeker, 1855) and Upeneus pori Ben-Tuvia & Golani, 1989
The goatfishes Upeneus moluccensis and U. pori are two Lessepsian fishes widespread in the
eastern Mediterranean (Zenetos et al. 2010). U. moluccensis is an old fish invader of the
Mediterranean, firstly recorded during the 1930’s in Israel (Gruvel 1931), while the congeneric
Upeneus pori, was reported a few years later from Turkey (Kosswig 1950). There have been
indications of competition for space and food between the two Upeneus species with the native
mullids Mullus spp. (Otero et al. 2013; Fanelli et al. 2015; Arndt et al. 2018). Nevertheless,
Upeneus spp. have become commercially important species for the fisheries of the eastern
Mediterranean (Keskin et al. 2011; Otero et al. 2013; Çiçek et al. 2014; Yemisken et al. 2014;
Turan et al. 2016; Gökçe et al. 2016a; Katsanevakis et al. 2018; Ali 2018; van Rijn et al. 2020).
54
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