Are Urban Mangroves Emerging Hotspots of Non-Indigenous Species? A Study on the Dynamics of Macrobenthic Fouling Communities in Fringing Red Mangrove Prop Roots

Abstract

Urbanization represent a radical transformation of natural habitats that alters all the biotic and abiotic properties governing ecosystems. Urban expansion often results in oversimplified communities, where most specialists decline or disappear and a few generalist or exotic species become dominant. The consequences of urban expansion in mangrove forests are understudied, although these systems have been altered by humans through centuries and the growth of human population in tropical coasts is expected to be faster than in higher latitudes. To assess the importance of indigenous and non-indigenous species in driving temporal and spatial changes in community structure of red-mangrove prop-root macrobenthic communities, we studied heavily altered mangrove forests from two bays from the Caribbean coast of Colombia in 2005 and 2021. In all places/periods, the community richness was low, a few taxa were dominant (11 taxa, out of 40, comprised ~ 90% of the total abundance) and the majority of all taxa (65%) were non-indigenous species whose presence is related with known stressors in urbanized systems. Hence, we suggest that urban mangrove forests are emerging hotspots for non-indigenous biota. Community structure did not change within or between bays, there was a clear, significant turnover of core species between 2005 and 2021, with non-indigenous species playing a prominent role in this variability. This was puzzling –ecological theory asserts that the abundance of a species is related to their permanence: core species are relatively stable through time while rare species appear or disappear– but this may not apply for communities dominated by non-indigenous biota.

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Are Urban Mangroves Emerging Hotspots of Non-
Indigenous Species? A Study on the Dynamics of
Macrobenthic Fouling Communities in Fringing Red
Mangrove Prop Roots
Enis Mosquera
Universidad de Antioquia
Juan Blanco-Libreros
Universidad de Antioquia
José M. Riascos (  josemar.rv@gmail.com )
Universidad de Antioquia
Research Article
Keywords: urban ecology, urban expansion, invasive alien species, biotic homogenization, Rhizophora
mangle.
Posted Date: March 21st, 2022
DOI: https://doi.org/10.21203/rs.3.rs-1440305/v1
License:   This work is licensed under a Creative Commons Attribution 4.0 International License. 
Read Full License

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Abstract
Urbanization represent a radical transformation of natural habitats that alters all the biotic and abiotic
properties governing ecosystems. Urban expansion often results in oversimplied communities, where
most specialists decline or disappear and a few generalist or exotic species become dominant. The
consequences of urban expansion in mangrove forests are understudied, although these systems have
been altered by humans through centuries and the growth of human population in tropical coasts is
expected to be faster than in higher latitudes. To assess the importance of indigenous and non-
indigenous species in driving temporal and spatial changes in community structure of red-mangrove
prop-root macrobenthic communities, we studied heavily altered mangrove forests from two bays from
the Caribbean coast of Colombia in 2005 and 2021. In all places/periods, the community richness was
low, a few taxa were dominant (11 taxa, out of 40, comprised ~ 90% of the total abundance) and the
majority of all taxa (65%) were non-indigenous species whose presence is related with known stressors in
urbanized systems. Hence, we suggest that urban mangrove forests are emerging hotspots for non-
indigenous biota. Community structure did not change within or between bays, there was a clear,
signicant turnover of core species between 2005 and 2021, with non-indigenous species playing a
prominent role in this variability. This was puzzling –ecological theory asserts that the abundance of a
species is related to their permanence: core species are relatively stable through time while rare species
appear or disappear– but this may not apply for communities dominated by non-indigenous biota.
1. Introduction
Habitat loss and invasive species have long been heralded as major causes of biodiversity loss in
conservation biology (e.g., Pimm & Raven 2000; Sala et al. 2000; Molnar et al. 2008). Among the vast
array of human transformations of natural habitats, cities represent the heart of our human enterprise
and perhaps the most radical source of ecological disturbances: urbanization alters all the abiotic and
biotic properties that govern ecosystems (Alberti 2008). Recent evidence suggest that the multifarious
human pressures clustered in cities is creating oversimplied communities where specialized species
decline and generalist tolerant species prevail (Faeth et al. 2011; Santana et al. 2020). This, and the
transport of people and goods —a major vector for species translocation—interact to foster cities as
emerging hotspots for the arrival, establishment and expansion of non-native species (Santana et al.
2020; Gonzalez-Lagos et al. 2021).
In urban ecology, much of the existing methods, principles, frameworks and knowledge come from
developed countries in the so-called Global North (Shackleton et al. 2021), where studies have been
traditionally focused in terrestrial socio-ecological systems. Fixing this is imbalance is critical for two
main reasons. First, major biodiversity hotspots in pan-tropical areas from developing countries are
forecasted to experience the fastest rates of urbanization by 2030, with consequences for biodiversity
loss (Seto et al. 2012). Second, coastal or riverine ports near coasts account for a considerable number of
the major cities of the Global South: 32 of the 77 largest cities in the world are located on coastal areas
of the Global South (Myers 2021). Third, population density is much higher in coastal areas (Faulkner

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2004) and compositional reorganizations related to human pressures are faster and more variable in
marine than in terrestrial ecosystems (Blowes et al. 2019).
Through centuries, mangrove forests have provided a multitude of ecosystem services and signicant
aesthetic, educational, cultural, recreational and spiritual benets for humans –more than any other
tropical coastal ecosystem (Millennium Ecosystem Assessment 2005). Yet for the same reason,
mangroves are highly threatened systems: mangrove forests are being lost at a fast rate in recent
decades, driven primarily by aquaculture development, deforestation, freshwater diversion and species
introductions commonly associated with urban expansion (Duke et al. 2007; Chakraborty 2019; Branoff
2017). Therefore, mangrove forests in urban settings are ideal systems to assess how indigenous and
non-indigenous biotas are interacting through time. Particularly, mangrove-root fouling communities are
suitable model systems to study the dynamics of community structure related to environmental changes
at different spatial scales, ranging from individual roots to whole mangrove islands (e.g. Farnsworth and
Ellison 1996, Hay et al. 2004).
In this study we assessed the relative importance of native and non-native species in driving temporal
and spatial changes in community structure of mangrove-root associated biota in the Urabá Gulf,
southern Caribbean coast of Colombia. Mangrove forests in this area have been heavily altered by land
reclamation for agricultural expansion in El Uno bay and urban expansion of the Turbo city port (Blanco
and Estrada-Urrea 2015). Owing to the intensication and heterogeneity of the anthropogenic impacts in
these areas during the last 15 years, we hypothesized that the structure of epibenthic macrofauna
associated with the roots of
R. mangle
will differ between periods (2005–2021) and bays (El Uno –
Turbo), with invasive species being important drivers of those differences. We built on previous historical
work performed by García and Palacios (2008) in the study area and predicted that the number of alien
and alien-invasive species will be consistently higher through time in Turbo bay, reecting its urban
character and thus a more diverse arrange of human activities and introduction vectors in this bay.
2. Materials And Methods
Study area
The study was performed in the southeastern coast of the Urabá Gulf (Fig.1). Located near the
Colombia-Panama border, the gulf is a north-facing embayment that represents the southernmost region
of the Caribbean Sea. The gulf is home to the most developed mangrove forests in the Colombian
Caribbean, which are probably the most productive in the Americas (Riascos and Blanco-Libreros 2019).
Fringe forest is the dominant physiographic type of mangroves in the region, which mostly comprise
monospecic stands of
Rhizophora mangle
while
R. mangle, Laguncularia racemosa
and
Avicennia
germinans
occur in basing mangroves (Urrego et al. 2014). The gulf is part of the Chocó-Darien Global
Ecoregion, a globally recognized biodiversity hotspots prioritized for conservation due to the high levels
of biodiversity and endemism (Fagua and Ramsey 2019). Despite this, the coalescence of outrages and
conicts that characterized the aftermath of European invasion in Latin America is epitomized in this

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region. The region witnessed the rose and dead of the oldest funded Spanish city of the Americas in
terra
rma
(Sarcina, 2017), the spread of African-descendant peoples that displaced indigenous groups after
slavery abolition and the growth of coca cultivation and the linked problem of illegal armed groups that
turned the region into a major tracking illegal immigration corridor. These historic processes have
resulted in a region displaying a complex mosaic of land covers, ethnic groups and legal and illegal
economic activities.
Mangroves in Turbo Bay have been characterized as “peri-urban” because they are structurally and
functionally affected by its proximity to the Turbo port city (Blanco-Libreros and Estrada-Urrea 2015).
During the last 15 years, the number of homes in the Turbo District increased by 18.12%, most of them
concentrated in Turbo city that currently is home to 48,787 people (DANE, 2018). Moreover, the ongoing
development of major port facilities will further boost urban expansion and the associated pressures on
mangrove forests in coming years. In turn, mangroves at El Uno Bay have been cleared for expanding
lands for agriculture (mainly comprising plantain crops) and cattle ranching, a typical example of a rural-
agricultural transition (Blanco-Libreros and Estrada-Urrea 2015). The bay is a coastal lagoon whose
formation is linked to the evolution of the Turbo River delta since the transfer of its mouth to this region in
the mid-20th century (Blanco-Libreros et al. 2013, Alcántara-Carrió et al. 2019).
Field work
This work builds on a previous work on the structure of macrobenthic communities associated to prop
roots of
R. mangle
performed by García and Correa (2006), which was latter published by García and
Palacio (2008). They sampled six prop roots in the eastern, western and northern zones of each bay
between September and December 2005. They found that diversity of macrobenthic communities did not
signicantly change trough time or zones. Moreover 12 species comprised 90% of the total abundance
and these species were found in all sampling points through the study period. Hence, we performed a
single sampling in June 2021, taking ten
R. mangle
prop roots in the same zones (east, north, west) in
each bay (Fig.1). Following the criteria established by García and Palacio (2008), roots were selected by
i) belonging to mature trees (≥ 10cm in diameter at breast height), ii) having a signicant portion
submerged into the water and iii) harboring easily seen sessile organisms. The roots were cut at the high-
tide mark and immediately stored in labeled plastic bags. Additionally, the following factors related to
anthropogenic disturbances were registered: trampling (the number of human footprints), logging
(number of trees cut), litter (number of litter items) and urban structures (number of urban structures, i.e.
houses, roads, peers, etc.). All these counts were performed by a single dedicated observer in the area
surrounding each sampling point.
Samples were immediately taken to the Marine Ecology laboratory (Universidad de Antioquia, Marine
Science Campus in Turbo), refrigerated at 3–5°C and processed within the next 12 hours. Roots were
weighted and placed on plastic trays, cut into small parts and dissected. Observed macroinvertebrates
were removed and stored in alcohol. Oysters in particular were carefully reviewed under a stereoscope to
record attached organisms. Plastic bags and root pieces were washed and sieved through a 250-micron

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mesh sieve. The retained material was stored in labeled plastic jars with 95% ethanol for further analysis.
The samples were sorted under a stereomicroscope and the resulting organisms identied to the
minimum possible taxonomic level. Following the criteria and denitions used by the Convention on
Biological Diversity on invasive alien species (https://www.cbd.int/invasive/terms.shtml), all taxa
identied to the species level were categorized as:
1. Indigenous species: a species living within its natural range (past or present) including the area
which it can reach and occupy using its natural dispersal systems.
2. Alien species: a species introduced outside its natural past or present distribution.
3. Invasive-alien species: an alien species whose introduction and/or spread threaten biological
diversity
Finally, a species that was not demonstrably native or introduced based on current knowledge was
classied as a cryptic species.
Data analysis
As sampled prop roots had distinct weight, the abundance of species was calculated as the number of
individuals per gram of root. These data were organized in biological (species abundances per root) or
environmental (anthropogenic disturbances in each sampling point) matrices. Abundance data were
square-root transformed to balance the contribution of abundant and rare species in further analysis,
thus accounting for the fact that some fast-moving animals had a chance to escape during samplings,
as opposite of sessile animals. Data on environmental factors were rst normalized (subtracting the
mean and dividing by the standard deviation for each variable) to account for the different scales among
variables. The Bray-Curtis dissimilarity index was later estimated from abundance data for each pair of
samples in the matrix and Non-metric multidimensional scaling (nMDS; Clarke and Gorley 2006) was
used to build ordination plots of the structure of macrobenthic communities in mangrove roots for each
zone and bay. In turn, Euclidean distances were calculated between sampling points to describe abiotic
differences among zones in each bay, using bi-dimensional plots of Principal Component Analyses.
To test for changes in the structure of epibenthic macrofauna associated to roots between zone (east,
north, west) and bays (El Uno, Turbo) we used a two-way ANOSIM test. This approach performs a
permutation test of the null hypothesis of no differences among a priori dened groups of samples,
based on the ranks of the sample dissimilarity matrix (Somereld et al. 2021). This preliminary analysis
conrmed that there were no signicant differences between zones.
To assess our hypothesis on changes in the structure of epibenthic macrofauna associated to roots,
samples from each zone were pooled and treated as replicates. A crossed two-way ANOSIM test was
used to test for differences between periods (2005–2021) and bays (El Uno – Turbo). For samples found
to be signicantly different, the Similarity Percentage Analysis (SIMPER) implemented in PRIMER
software was used to evaluate which species contributed most to the differences between periods and
bays. This biota was further characterized according to size and origin (native/non indigenous) to

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discuss our ndings. A signicance level of α = 0.05 was chosen for all the tests performed. All
multivariate analyses were performed
using PRIMER v.6 software (Clarke and Gorley 2006).
3. Results
Our study reports 40 taxa of invertebrates associated to prop roots of
Rhizophora mangle
(Table1) in El
Uno and Turbo. We found 225 specimens in 2021 (El Uno = 75; Turbo = 150), which was nearly half of the
394 specimens found in 2005 (El Uno = 201; Turbo = 193). Richness (number of taxa) also decreased in
2021 (El Uno = 20; Turbo = 18) with respect to 2005 (El Uno = 28; Turbo = 29). A few taxa where highly
dominant in all places/times: 11 species comprised more than 90% of the abundance. In fact, two
species (
Brachidontes dominguensis
and
Tanais dulongii
) made up nearly half of the total abundance. In
contrast, there were 28 taxa that contributed less than 1% of the total abundance. Of the 40 taxa found in
our study, 26 (65%) were identied to the species level and categorized as indigenous (12), alien (7),
invasive alien (5) and cryptogenic (2) (Table S1).

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Table 1
Abundance (expressed as percentage of total abundance shown in the last row) of invertebrate species
associated to prop roots of
Rhizophora mangle
in 2005 and 2021 at El Uno bay and Turbo bay,
Colombian Caribbean. Note that total abundances for 2005 at the two bays represent the average for the
period September to December.
Percentage of total abundance 
 2005 2021  
Taxa El Uno Turbo El Uno Turbo Total Cumulative
Brachidontes domingensis
(Lamarck. 1819) 33.762 38.844 7.368 10.979 26.555 26.555
Tanais dulongii
(Audouin. 1826) 23.138 20.255 13.158 19.643 20.169 46.725
Crassostrea rhizophorae
(Guilding.
1828) 13.571 16.533 2.105 2.447 10.375 57.100
Leptocheirus rhizophorae
(Ortíz &
Lalana. 1980) 0.000 0.000 27.237 26.257 9.734 66.834
Apocorophium acutum
(Chevreux.
1908) 1.930 0.130 25.658 16.138 7.739 74.573
Mytilopsis sallei
(Récluz. 1849) 3.030 4.869 3.553 8.598 5.021 79.594
Exaiptasia diaphana
(Rapp. 1829) 11.724 1.320 0.000 0.000 4.203 83.798
Biustra tenuis
(Desor. 1848) 0.000 0.022 6.447 7.341 2.585 86.383
Alitta succinea
(Leuckart. 1847) 1.556 2.683 0.921 1.786 1.884 88.267
Sphaeroma terebrans
(Bate. 1866) 1.619 0.887 5.395 0.992 1.701 89.968
Amphibalanus amphitrite
(Darwin.
1854) 2.200 2.380 0.395 0.331 1.579 91.548
Panopeus herbstii
H. Milne
Edwards. 1834 1.473 3.008 0.132 0.000 1.426 92.974
Bankia mbriatula
(Moll & Roch.
1931) 0.062 2.813 0.000 0.000 0.893 93.867
Aratus pisonii
(H. Milne Edwards.
1837) 1.349 0.692 0.658 0.463 0.845 94.712
Vitta virginea
(Linnaeus. 1758) 0.083 2.337 0.395 0.000 0.800 95.512
Chaetopteridae (Audouin & Milne
Edwards. 1833) 0.000 0.000 4.737 0.860 0.790 96.302
Brachidontes
(Swainson. 1840) 2.137 0.000 0.000 0.000 0.692 96.993
Sabellidae (Latreile.1825) 0.000 0.000 1.053 1.720 0.548 97.541

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Percentage of total abundance 
Thaisella coronata
(Lamarck.
1816) 0.166 0.216 0.263 0.860 0.363 97.904
Littoraria angulifera
(Lamarck.
1822) 0.498 0.433 0.000 0.000 0.295 98.199
Chthamalus
(Ranzani. 1817) 0.519 0.390 0.000 0.000 0.289 98.488
Nereis
(Linnaeus. 1758) 0.623 0.130 0.000 0.000 0.242 98.730
Pachygrapsus gracilis
(de
Saussure. 1857) 0.042 0.714 0.000 0.000 0.235 98.965
Neoteredo reynei
(Bartsch. 1920) 0.000 0.260 0.000 0.529 0.209 99.174
Ligia
(Fabricius. 1798) 0.042 0.606 0.000 0.000 0.201 99.376
Ascidia (Linnaeus. 1767) 0.000 0.022 0.000 0.661 0.168 99.543
Neopanope
A. Milne-Edwards.
1880 [in A. Milne-Edwards. 1873–
1880]
0.042 0.303 0.000 0.000 0.107 99.651
Hirudinea (Savigny. 1822) 0.000 0.022 0.000 0.265 0.071 99.722
Polymesoda arctata
(Deshayes.
1855) 0.145 0.022 0.000 0.000 0.054 99.776
Pyrgophorus
(Ancey. 1888) 0.104 0.000 0.000 0.000 0.034 99.809
Stenoninereis
(Wesenberg-Lund.
1958) 0.000 0.000 0.000 0.132 0.032 99.842
Martesia striata
(Linnaeus. 1758) 0.000 0.043 0.132 0.000 0.030 99.871
Diptera (Linnaeus. 1758) 0.000 0.022 0.132 0.000 0.023 99.894
Sphaeroma
(Bosc. 1801) 0.062 0.000 0.000 0.000 0.020 99.914
Culex pipiens
(Linneus. 1758) 0.000 0.000 0.132 0.000 0.016 99.930
Platynereis mucronata
(León-
González. Solís-Weiss & Valadez-
Rocha. 2001
0.000 0.000 0.132 0.000 0.016 99.946
Pyrgophorus parvulus
(Guilding.
1828) 0.042 0.000 0.000 0.000 0.013 99.960
Macrobrachium acanthurus
(Wiegmann. 1836) 0.021 0.022 0.000 0.000 0.013 99.973
Callinectes sapidus
(Rathbun.
1896) 0.021 0.022 0.000 0.000 0.013 99.987

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Percentage of total abundance 
Callinectes bocourti
A. Milne-
Edwards. 1879 [in A. Milne-
Edwards. 1873–1880]
0.042 0.000 0.000 0.000 0.013 100
Richness (number of taxa) 28 29 20 18 40 
Total abundance (Number of
individuals-all taxa) 201 193 76 151 621 
The abundance and species composition of macrobenthic assemblages associated to roots of
R. mangle
did not show signicant differences among zones, neither in El Uno bay (ANOSIM; R = 0.012;
p
= 0.34) nor
in El Uno Bay (ANOSIM; R = 0.116;
p
= 0.06), although a higher spatial segregation in nMDS ordination
plots was observed in Turbo bay (Fig.2). This result was surprising, because PCA ordinations of
anthropogenic pressures (Fig.3) showed a clear spatial structure within each bay, with the density of litter
and urban structures being key structuring factors.
In contrasts, our results rendered signicant differences in the structure of benthic assemblages between
periods (ANOSIM; R = 0.853; p = 0.02) but not between bays (ANOSIM; R = 0.000; p = 0.45). These results
are also illustrated in the nMDS ordination plot (Fig.4), which show that samples from 2005 clustered to
the left of the plot while samples from 2021 clustered to the right. In turn, between-bays distances were
less consistent. Results of SIMPER analysis showed that 12 species explained more than 90% of
dissimilarity in abundance and species composition between samples taken in 2005 and 2021 (Table2).
Of these species, only ve were categorized as indigenous species, while the majority where either, alien,
invasive alien or cryptogenic species (Fig.5; Table S1). Of special importance was the fact that two of the
most abundant species observed in 2005 where replaced as dominant species in 2020 by two previously
unregistered species.

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Table 2
Results of Similarity Percentage Analysis showing the main benthic species contributing to the
dissimilarity in community structure between 2005 and 2021
Main discriminating species Abundance average Dissimilarity
Average
Contribution
(%)
Cumulative (%)
 2005 2021   
Branchidontes domingensis
23.76 3.70 18.6 24.96 24.96
Leptocheirus rhizophorae
0.00 10.07 9.96 13.36 38.31
Tanais dulongii
14.24 6.62 9.46 12.69 51.00
Crassostrea rhizophorae
9.85 0.88 8.46 11.34 62.35
Apocorophium acutum
0.69 7.32 6.87 9.21 71.56
Exaiptasia pallida
4.35 0.00 4.72 6.34 77.90
Biustra tenuis
0.01 2.67 2.53 3.40 81.30
Mytilopsis sallei
2.58 2.62 2.17 2.91 84.20
Panopeus herbstii
1.46 0.02 1.38 1.85 86.05
Vitta virginea
0.78 0.05 1.31 1.76 87.81
Amphibalanus amphitrite
1.50 0.13 1.26 1.69 89.50
Alitta succinea
1.38 0.57 1.09 1.47 90.97
4. Discussion
In a classical work on ecosystem ecology, Odum (1985) suggested that in a system submitted to external
disturbances or stressors, species diversity would decrease while dominance increase. Direct,
comparisons of species diversity of fouling communities associated to roots between different places are
dicult to make because of differences in sampling effort, spatial coverage, taxonomic expertise, abiotic
conditions, large scale trends in biodiversity, among others. Despite this, the richness of taxa in our work
was much lower than that typically found in non-urban spots of the Caribbean Sea, including Belize
(Ellison and Farnsworth 1992 = 46 taxa; Farnsworth and Ellison 1996 = 59 taxa), Mexico (Hemández-
Alcántara and Solís-Weiss 1995 = 86 taxa; Tunell and Withers 1996 = 47–56 taxa; Lucas and de la Cruz-
Francisco 2018 = 26 taxa; Ruiz and López-Portillo 2014 = 28 taxa), Venezuela (Guerra-Castro et al. 2011 =
115 taxa that included algae). Because of this, and the decreasing richness observed between in 2021
compared to 2005, we suggest that the low richness of taxa of fouling communities in red mangrove
roots might be a response to impacts of increased urbanization in Turbo and related human activities
that spill-over in El Uno. More importantly, this is in line with the observed variation of faunal and plant

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species richness along rural-urban gradients (Alberti 2008). However, a second explanation for the low
richness may also lie in the fact that estuarine conditions in the study area restrict stenohaline species.
In contrast, we could not suggest a link between the high dominance of a few species and disturbances
associated to urbanization –remarkably high dominances in fouling communities or specic taxonomic
assemblages in mangrove roots have been observed in relatively undisturbed areas (Hemández-Alcántara
and Solís-Weiss 1995; Farnsworth and Ellison 1996; Tunell and Withers 1996; Vilardy and Polania 2002;
Molina et al. 2017).
Mangrove-root epibenthic communities have long been recognized as spatially structured communities
controlled by physical and biological factors (Binham 1992; Farnsworth and Ellison 1996). At local
scales, changes in community structure are known to be mainly controlled by larval supply: the
patchiness seen in many mangrove epifaunal communities is largely a result of the importance of short-
lived lecithotrophic species (e.g., sponges, bryozoans, ascidians), while homogeneous communities result
from the dominance of species with long-lived planktotrophic larvae (Binham 1992). Thus, given the lack
of spatial differences in the structure of mangrove-root benthic communities in our study we would
expect that species with long-lived planktotrophic larvae are dominant. But It seems not the case: for
example,
Tanais dulongii
(which alone comprised 20% of the total abundance), is a brooding crustacean
with a strictly benthic life cycle and low dispersion rate (Rumboldt et al. 2015).
The lack of spatial variability in community structure seems more likely related to i) the surprising nding
that non-indigenous species outnumber indigenous species and ii) the local expression of a widely
observed effect of species invasions: biotic homogenization (
sensu
McKinney and Lockwood 1999).
Signicant increases in the number of non-indigenous plant species as a response to urbanization have
been observed in forested urban wetlands (Ehrenfeld and Schneider 1991, Paquin et al. 2021) and urban
mangroves (Branoff and Martuzzi 2020). In our knowledge, this is the rst study showing that animal
assemblages associated to mangrove forests in urban areas are dominated by non-indigenous species;
hence, we suggest that urban mangroves may be emerging hotspots for non-indigenous biota. Some of
the most abundant species in our study are invasive species associated to conditions commonly found in
urbanized coasts:
Balanus Amphitrite, Mytilopsis sallei
and
Alitta succinea
are common elements of
encrusting communities in human-made structures worldwide or in invasive mangroves (Neves and
Rocha 2008; Demopoulos and Smith 2010; Aguilera et al. 2018; Tan Tay 2018, );
Apocorophium acutum
is an invasive species associated algal mats in jetties and aquaculture facilities (Hossain and Hughes
2016, (Giménez-Delcamp 2021);
Exaiptasia diaphana
is an alien species covering articial and natural
surfaces (Durán-Fuentes 2022) and
Tanais dulongii
is a cryptogenic species commonly found in
eutrophic waters (Wildsmith et al. 2009).
A second striking results in our study was a clear turnover of core species between 2005 and 2021, with
non-indigenous species playing a prominent role in this variability. A fairly common feature of ecological
communities is that a few (core) species are exceptionally abundant, whereas most are rare, transient
species (Gaston and Blackburn, 2000). Empirical evidence show that core species tend to be present for a

Page 12/21
longer period than rare species, thus implying that the commonness and rarity of species in the
assemblage is related to their permanence (e.g. Magurran & Henderson 2003). Therefore, the observed
turnover in core species is unexpected and hard to explain within the scope of our data. As invasion is not
an event but a species-specic and site-specic process occurring in consecutive stages including
transport, establishment, spread and impact (Lockwood et al. 2013). Thus, we hypothesize that the
observed turnover may reect the progress or failure of each species to go through these stages. This,
however is rather speculative because our data have intrinsic limitations mostly related with the fact that
we comparing start and end conditions without knowledge of e.g. long-term or cyclic environmental or
oceanographic changes between these conditions that may be independent of urban expansion.
The unusually high proportion of non-indigenous species in mangrove-root epibenthic communities and
the dicult to understand the signicant species turnover of core species most probably reect potential
synergistic effects of multiple co-occurring stressors on the establishment and impact of non-native
species—a current research priority to advance invasion science in the face of rapid environmental
change (Ricciardi et al. 2021)
Declarations
Author Contributions
All authors contributed to the study conception and design. Field work was conducted by José M.
Riascos and Enis Mosquera. Material preparation, data collection and analysis were performed by José
M. Riascos and Enis Mosquera. The rst draft of the manuscript was written by José M. Riascos and Enis
Mosquera, and all authors commented on previous versions of the manuscript.
Funding
This work was nancially supported by a ICETEX scholarship for black minorities and Esperanza Afro
foundation, awarded to Enis Mosquera.
Acknowledgements
María José Pacheco contributed to eldwork. Thanks to local communities in El Uno and Turbo bay for
allowing access to their territories. Fieldwork and collection of biological samples was conducted in
under the permit 0524 provided by Autoridad Nacional de LicenciasAmbientales to Universidad de
Antioquia.
Competing interest
The authors declare that the research was conducted in the absence of any commercial or nancial
relationships that could be construed as a potential conict of interest.
Data Availability

Page 13/21
All data generated or analysed during this study are included in this published article [and its
supplementary information les]. Any further information regarding the database supporting this work is
available from the corresponding author on reasonable request.
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Figures

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Figure 1
Map of the Urabá gulf and the study sites. Shape of mangrove forests provided by Valencia-Palacios and
Blanco-Libreros (2021)
Figure 2
Ordination by non-metrical Multidimensional Scaling of composition and abundance of macrobenthic
communities associated to prop roots of
Rhizophora mangle
in El Uno Bay (a) and Turbo bay (b),
Colombian Caribbean coast. Ordination maps were calculated from Bray-Curtis dissimilarity measures.
Figure 3

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Bi-dimensional plots of Principal Components (PC1 and PC2) after the Principal Component Analyses of
anthropogenic factors registered in sampling spots in western, northern and eastern zones at El Uno Bay
(a) and Turbo Bay (b) and the superimposed vectors (grey lines) of anthropogenic factors.
Figure 4
Ordination by non-metrical Multidimensional Scaling (nMDS) of composition and abundance of
macrobenthic communities associated to prop roots of
Rhizophora mangle
in El Uno Bay and Turbo bay,
Colombian Caribbean coast. nMDS was built on Bray-Curtis dissimilarity measures.

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Figure 5
Abundance of Indigenous, alien, alien-invasive and cryptic macrobenthic species mainly contributing to
the dissimilarity in community structure between 2005 and 2021 (after SIMPER analysis) in Turbo and El
Uno. Note that, together these species comprised about 90% of the total abundance in each study period.
References for the categorization of this, and the full set of species found in this study, are given in Table
S1; note that
Crassostrea rhizophorae
and
Mytilopsissallei
are native to the Caribbean, but they have
been introduced to other regions and hence are reported here as alien and invasive-alien species,
respectively.
Supplementary Files
This is a list of supplementary les associated with this preprint. Click to download.
TableS1.xlsx