1 Introduction

Urban coastal wetlands are fragile, highly productive ecosystems that provide important ecosystemic services (Clarkson et al., 2014). They regulate nutrient cycles and water quality by retaining, transforming, or removing nutrients and pollutants (Neiff, 1999; Novoa et al., 2020), and help to regulate aquifers (Goldsborough and Robinson, 1996) by promoting an excess of organic matter and high primary plant productivity that help support trophic chains (Elser and Frees, 1995). However, urban coastal wetlands are disappearing as a result of the extensive urban growth, a process that causes serious alterations in their structure, functioning, and diversity due to the advancement of urban, agricultural, foraging, and drainage activities, and the displacement of endemic species by invasive ones (Li et al., 2018; Albert et al., 2021).

Urban development has become a dominant force defining ecosystem patterns and processes all over the world (Vitousek et al., 1997; Alberti et al., 2003). It has been shown that the level of urbanization correlates with reductions in the number of species – the greater the urban development in an ecosystem, the lower the species richness and the higher the functional homogeneity of the endemic biota (Devictor et al., 2007; Buchholz et al., 2018; Piano et al., 2020). These ecosystemic changes are likely to effect different inter- and intraspecific relationships. Despite their influence, the impact of changes in soil coverage typically associated with urbanization is seldom included in analyses of the trophic ecology of the impacted ecosystem (Wu and Shaner, 2016).

Arthropods are play an important role in urban coastal wetlands. For example, it has been documented that birds – widely represented in these ecosystems – as well as amphibians and some insectivore mammals exploit the proliferation of introduced arthropods as food resource (Hooks et al., 2003). Also, pollination of the wetland’s native flora is accomplished mainly by some insects and also by some arachnids, which other in addition to the previously mentioned factors, shows the trophic and ecological significance of arthropods in these fragile ecosystems (Correa-Araneda et al., 2011; Novoa et al., 2020). However, studies on the relationships of arthropod assemblages in urban ecosystems, particularly urban coastal wetlands, are few and far between (Castillo-Velásquez and Huamantinco-Araujo, 2020).

In northern Chile, arthropods are one of the most abundant and diverse biological taxons (Cepeda-Pizarro et al., 2005a; Pizarro-Araya et al., 2008) and have multiple ecological roles. During the humid phase, they are important pollinators that help dynamize energy flow and nutrient cycles, and are an abundant high-quality trophic resource (Cepeda-Pizarro et al., 2015, 2016); during the arid phase (dry periods), they are important macrodecomposers and an abundant trophic resource for vertebrates (Torres-Contreras et al., 1994; Torres-Contreras, 2001; Pizarro-Araya, 2010).

Knowledge of arthropods inhabiting coastal wetlands in Chile is limited to studies conducted in marsh urban wetlands of central-southern Chile, such as Sepúlveda-Zuñiga et al. (2012), who assessed the vegetation naturalness and heterogeneity levels and their impact on Macrolepidoptera diversity (Insecta: Lepidoptera), and Villagrán-Mella et al. (2006), who studied the relationship between habitat features and insect assemblage structure. More recently, De los Ríos et al. (2019) and Zuleta et al. (2019) documented the aquatic arthropod fauna of ephemeral ponds in Huentelauquén (Coquimbo), which are habitats that host an arthropod community that differs from that found in Chilean continental waters. These studies have characterized species such as Lynceus huentelauquensis Sigvardt et al., 2019 (Crustacea: Branchiopoda), the first Lynceus record for Chile and the 13th for the American continent (Sigvardt et al., 2019). A molecular analysis placed L. huentelauquensis in a clade close to Australian species (Sigvardt et al., 2019).

In the northern part of the Chilean coastal desert, represented by the plant formation of the coastal desert of Tocopilla (Gajardo, 1993) and the Morro Moreno National Park (23° S) and La Chimba National Reserve (23° S), knowledge of the arthropod community is scarce and scattered. These units are expected to contain a unique endemic arthropod fauna, as documented for the coastal Paposo priority area (25° S), an area hosting numerous endemic arthropod species. As for coastal streams and their associated wetlands, knowledge is almost nonexistent and limited only to some published records of water snails of the genus Heleobia (Collado et al., 2019).

Therefore, the purpose of this study was to characterize the richness, abundance, and origin of the terrestrial arthropod fauna of the Aguada La Chimba urban coastal wetland, as well as current threats to this local fauna and the challenges for their preservation.

2 Material and methods

2.1 Study area

The Aguada La Chimba urban coastal wetland is located in the northern part of Antofagasta city. The area constitutes a different ecosystem within the arid matrix of the Antofagasta coast whose vegetation is supported by groundwater overflows from two streams: El Rubio (northern part) and Chimbanito (southern part) (Fig. 1). The Chilean Ministry of Environment declared the wetland a “Natural Sanctuary” and “Urban Wetland” with the main goal of preserving Heleobia chimbaensis (Biese, 1944) (Cochliopidae), a threatened snail from La Chimba (D.S. No. 52/2014 MMA) that is highly susceptible to pressures and threats due to its reduced habitat and high dependency on water availability (MMA, 2014).

Fig. 1
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Geographical location of the Aguada La Chimba urban coastal wetland (Antofagasta Region, Chile). Codes as in Fig. 2

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Fig. 2
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Study sites (habitats) in the Aguada La Chimba urban coastal wetland: a panoramic view of the wetland; b Chimbanito stream (CH2) with Arundo donax Linnaeus, 1753 (Poaceae); c Chimbanito stream (CH1, CH3) with Tessaria absinthioides (Hook. & Arn.) DC., 1836 (Asteraceae); d El Rubio stream (RU) in the northernmost part of the wetland; e southernmost part of the wetland (RO)

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2.2 Arthropod sampling and identification methods

We selected 5 study sites with varying soil and vegetation characteristics representing different environments of the coastal wetland (Fig. 2). In each site, we set up 10 pitfall traps, a method that is widely used to assess epigean arthropod assemblages (Cepeda-Pizarro et al., 2005a, 2005b; Pizarro-Araya et al., 2014a, 2014b). The traps were installed 1 m apart from each other among representative vegetation of each environment sensu Chely and Corley (2010). Each trap consisted of two plastic cups one inside the other, with the interior cup easily removable (sensu Cepeda-Pizarro et al., 2005a, 2005b). The cups measured 7.4 and 7.6 cm in diameter by 10.2 and 12.0 cm in height, respectively. The interior cup was two-thirds filled with a solution of water, domestic washing liquid, and 70% alcohol. The traps remained active between December 17 and December 20, 2020.

In addition to pitfall trap samplings, we also conducted manual collections in those same environments. The collections were carried out by 3 collectors in linear transects 200 m in length for around 30 min (sensu Caldas and Robbins, 2003). The collectors concentrated in a strip 2 m wide by 3 m high along the transect, using an entomological net to collect flying species (Coleoptera, Lepidoptera, Hymenoptera, Diptera). These additional sampling efforts were repeated during the same days described above.

Scorpions (Scorpiones) were sampled using the previously described manual collection method with LED flashlights equipped with ultraviolet light (138 LED-UV) and powered by a 12 V battery.

A few specific nocturnal flying arthropod taxa (e.g., Coleoptera, Lepidoptera, Hymenoptera, Neuroptera) were sampled using the white-light method, which consists in setting up an artificial light source behind a horizontally placed white fabric sheet or cloth to attract individuals. Four light points behind two white sheets were used as long-distance attractants (Sheikh et al., 2016).

All the captured specimens were removed, cleaned, and preserved in alcohol (e.g., insects in 70% alcohol; crustaceans in 75% alcohol; arachnids in 80% alcohol) until their processing and taxonomic identification. The captured specimens are currently deposited at the entomology and arachnid collection of the Laboratorio de Entomología Ecológica of Universidad de La Serena, Chile (LEULS).

2.3 Data analysis

The richness of the different arthropod groups was estimated using the non-parametric estimators ICE (incidence-based coverage estimator), Chao 1, Chao 2, and first- and second-order Jackknife (Jack 1 and Jack 2) (Chazdon et al., 1998; Colwell, 2013). These estimators are universally valid for any species abundance distribution and more robust than estimators based on parametric models of species abundance (Chao and Chiu, 2016). For this reason, they provide more precise estimations of species abundance (Hortal et al., 2006). ICE is a robust and precise estimator of species richness (Chazdon et al., 1998), whereas Chao 1 and 2 and Jack 1 and 2 are based on rare species and provide less biased estimates for small samples (Colwell and Coddington, 1994). All the formulas of the estimators are found in Colwell (2013). All the estimators were calculated using EstimateS version 9.1.0 (Colwell, 2013). We calculated common indices of diversity for the studied sites based in the Hill series: species richness (S), the exponential Shannon–Wiener (exp H’), inverse of Simpson´s concentration index (1/D) and Berger-Parker index (1/d) (Magurran 2004). We compared the species diversity between studied sites by using rarefaction and extrapolation curves based on the first three Hill numbers (Chao et al., 2014) because these diversity metrics are interpreted as the "effective number of species" and they are not biased towards rare species (Jost, 2007). The confidence intervals for curves were obtained as mean of 100 bootstrap. Sample-based rarefaction/extrapolations were performed with iNext Online (Chao et al., 2016). Analysis of variance (ANOVA) was performed to assess differences between sites using each of the Hill numbers. Finally, to evaluate the rarity based on the abundance of species in the community, we compared species abundance distribution (SAD) models (geometric series, log-series and Poisson lognormal) with PAST 2.16 (Hammer et al., 2001).

2.4 Origin of the arthropod fauna

In order to identify the origin and current status of the species recorded in the study area, we conducted a literature review based on current knowledge and on the taxonomic resolution of the recorded taxa. We identified native, endemic, and naturalized species. In general, we considered as native species those that naturally inhabit in Chile; that is, those that are believed to have originated or have naturally arrived in the country without people intervention (MMA, 2021). Endemic species were those restricted to a specific territory, either a continent, country, political administrative unit, biogeographic region, island, or area (MMA, 2021). For this study, only endemic species that inhabit in Chile were included. Finally, naturalized species were those that are not endemic to Chile and were intentionally or accidentally introduced as a result of human activities (sensu Fuentes et al., 2020).

3 Results and discussion

3.1 Arthropod richness and abundance

All our sampling efforts yielded a total of 1.808 arthropod specimens. Although the UV-light method is a collection technique especially used for scorpions, no individuals from this group were captured. From the material collected from all the study sites, we identified a total of 111 arthropod species from 4 classes: Arachnida, Chilopoda, Crustacea, and Insecta. Insects (Insecta) were the most represented taxon, with 87 species, 47 families, and 15 orders. Arachnids (Arachnida) were represented by 19 species, 14 families, and 2 orders. Crustaceans (Crustacea) were represented only by 4 species, 4 families, and 2 orders. Finally, centipedes (Chilopoda) were represented by a single family. The most represented insect orders were Diptera, Hymenoptera, Coleoptera, and Lepidoptera, whereas Araneae was the richest group within Arachnida. Crustacea and Chilopoda were only represented by a few species (Table 1).

Table 1 Taxonomic composition of arthropods identified in the urban coastal wetland of Aguada La Chimba (Antofagasta Region, Chile)
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All the non-parametric estimators showed estimated richness values above the observed values for the different groups under study. Insects were the group with the highest estimated number of species in all indices (Table 2), which suggests that this taxon contains even more unrecorded species in the area. This is a common finding in studies of large taxonomic groups (e.g., arthropods) that use different sampling scales (Legros et al., 2019; Pryke and Samways, 2009; Andersen et al., 2010). Within insects, we found a significant number of species represented by single individuals (singletons) (Table 2). This may be explained, on one hand, by the existence of arthropod communities with rare species restricted to specific environments within the arid matrix or, on the other, by the imbalance in the dynamics of these communities (Coddington et al., 2009; Richardson and Arias-Bohart, 2011) resulting from the pressure exerted by predators and parasitoids on common species or from the grouped distribution of species. The latter may be dependent on the reproductive structure of arthropod species or the uneven distribution of resources (Longino et al., 2002). Rarefaction and extrapolation curves based on sample size suggest that the rate of species accumulation did not differ between sites which is evidenced by the overlap of the confidence intervals for diversity estimates when q = 0 and q = 1, except for CH2 when q = 2. On the other hand, the coverage based rarefaction and extrapolation curves showed the highest estimates of species richness (q = 0, q = 1, q = 2) for all sites studied except CH2 (Fig. 3). The ANOVA did not show significant differences between sites when we compared each of the Hill numbers; q = 0 (F4,5 = 0.138, p = 0.961), q = 1 (F4,5 = 0.177, p = 0.941), q = 2 (F4,5 = 0.273, p = 0.884) which suggests that the different sites of this wetland have a similar arthropod fauna associated with fragmented environments with a semiarid matrix more extense. These estimates of species richness suggest the need to conduct samplings in other seasons to improve the representativeness of this inventory.

Table 2 Estimated richness for arthropods identified in the urban coastal wetland of Aguada La Chimba (Antofagasta Region, Chile)
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Fig. 3
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Sample-and coverage-based rarefaction and extrapolation of Hill numbers for arthropods in different studied sites in the Aguada La Chimba urban coastal wetland: a Sample-sized-based rarefaction curve when q = 0; b Sample-coverage-based extrapolation curve when q = 0; c Sample-sized-based rarefaction curve when q = 1; d Sample-coverage-based extrapolation curve when q = 1; e Sample-sized-based rarefaction curve when q = 2; f Sample-coverage-based extrapolation curve when q = 2. Species richness (S) (q = 0), the exponential Shannon–Wiener (exp H´) (q = 1), the inverse Simpson index (1/D) (q = 2). The confidence intervals for curves were obtained as mean of 100 bootstrap

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Insects were the most abundant group, with 1,148 specimens (63.5%), followed by crustaceans, with 419 specimens (23.2%), and arachnids, with 235 specimens (13.0%). The most abundant insect orders were Diptera, Hymenoptera, and Collembola, whereas the most abundant arachnid order was Araneae (Table 1). The most abundant species were the amphipod Hyalella costera González and Watling, 2001 (Amphipoda: Hyalellidae) (13.07%), the isopod Benthanoides sp. (Isopoda: Philosciidae) (7.63%), and a single unidentified collembolan species of the family Poduridae (Collembola: Poduridae) (6.56%). These species were found in most environments of the Aguada La Chimba wetland, but they were more abundant in Chimbanito stream. The rank abundance curve showed low uniformity for the arthropod assemblage, with patterns of high dominance and rarity (Fig. 4). The Chimbanito (Chimbanito 1) and El Rubio streams were the sites more diverse and with less dominance, both with higher vegetation coverage (Table 3). The higher abundance observed in El Rubio stream was due to a few groups, including Myrmicinae (Hymenoptera: Formicidae), Entomobryidae (Collembola), and Delphacidae (Homoptera), the latter associated with monocotyledons (Wilson, 1997) such as Paspalum vaginatum Swartz, 1788 (Poaceae) and Distichlis spicata (I.) Greene (Poaceae), both present in this environment.

Fig. 4
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Species-abundance distribution and model fittings for arthropods from the Aguada La Chimba urban coastal wetland (Antofagasta Region, Chile): a Geometric model; b Log-series model; c Poisson lognormal model

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Table 3 Hill numbers for different studied sites in Aguada La Chimba urban coastal wetland (Antofagasta Region, Chile)
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3.2 Native, endemic, and naturalized arthropod species

Among the total number of species, the proportions of native (24%) and naturalized (22%) species were comparable, whereas the proportion of endemic elements was low. A significant proportion of species were not categorized due to the lack of available data in the literature (Fig. 5). One of the arachnids recorded in most study sites was the orb-weaver spider Argiope argentata (Fabricius, 1775) (Araneidae) (Fig. 6a), a native species widely distributed in South America and distributed in Chile from Arica to Antofagasta (Taucare-Ríos, 2012a). Another native South American spider recorded in all the study sites was Loxosceles laeta (Nicolet, 1849) (Sicariidae), a species of medical importance responsible for causing loxoscelism (Taucare-Ríos, 2012b).

Fig. 5
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Origin of the species recorded in the Aguada La Chimba urban coastal wetland (Antofagasta Region, Chile)

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Fig. 6
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Different species of arthropods recorded in the Aguada La Chimba urban coastal wetland: a female of Argiope argentata (Fabricius, 1775) (Araneae: Araneidae); b lateral view of Chileuma sp. (Araneae: Gnaphosidae); c lateral view of Lucilia sericata (Meigen, 1826) (Diptera: Calliphoridae); d lateral view of Hylephila fasciolata (Blanchard, 1852) (Lepidoptera: Hesperiidae); e lateral view of Pycnoscelus surinamensis (Linnaeus, 1758) (Blattodea: Blaberidae)

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Within Insecta, some globally distributed species recorded were the earwig Euborellia annulipes (Lucas, 1847) (Anisolabididae) and the cockroach Pycnoscelus surinamensis (Linnaeus, 1758) (Blaberidae) (Fig. 6e), the latter first recorded for continental Chile in 2018 (Taucare-Ríos et al., 2018). These species were recorded only in Paspalum vaginatum patches near El Rubio stream and are believed to be introduced species already established in national territory associated with gardens, city squares, and orchards (Arellano, 2014; Taucare-Ríos et al., 2018).

We also recorded some pest-control species, including the native coccinellid Eriopis chilensis Hofmann, 1970 (Coccinellidae), the introduced coccinellid Cryptolaemus montrouzieri Mulsant, 1853 (Coleoptera: Coccinellidae), and the introduced lacewing Chrysoperla externa externa (Hagen, 1861) (Chrysopidae). These insects feed on some species of agricultural importance, such as aphids (Aphididae), mealybugs (Pseudococcidae), and acarines (Acari) (Giffoni et al., 2007). These groups were recorded mostly in Chimbanito stream, a site characterized by the presence of Arundo donax Linnaeus, 1753 (Poaceae) and Tessaria absinthioides (Hook. & Arn.) DC., 1836 (Asteraceae).

Among the endemic elements, we recorded a spider of the genus Chileuma Platnick, Shadab and Sorkin, 2005 (Gnaphosidae) whose individuals may belong to Chileuma paposo Platnick, Shadab and Sorkin, 2005; this species is part of the epigean predatory fauna of these ecosystems (Fig. 6b). Other endemic species recorded were the amphipod Transorchestia chiliensis Milne-Edwards, 1840 (Talitridae), distributed from Antofagasta to Magallanes and associated with rocky intertidal zones (González, 1991); the butterflies Hylephila fasciolata (Blanchard, 1852) (Hesperiidae) (Fig. 6d) and Vanessa carye (Hübner, 1812) (Nymphalidae), both widely distributed in Chile, and the tenebrionid Scotobius tarapacensis Marcuzzi, 1976 (Coleoptera), a species that is rare in collections that was recorded in Tarapacá (Marcuzzi, 1976). In general terms, we observed an arthropod fauna made up of a fraction with native species with a more or less wide distribution in northern Chile, a fraction with naturalized species associated with some plant species, and a small fraction of endemic species associated with semi-arid ecosystems.

3.3 Current threats and local challenges

Wetlands are typically one of the most threatened ecosystems mainly due to the pressures from human activities and urban development. Consequently, these areas may receive residues (Boavida, 1999) or be subject to changes in their size, structure, hydrology, and biological communities as a result of climate change (IPCC, 2014; Barros and Albernaz, 2014; Salimi et al., 2021) or to the extraction of surface or underground water at basin level. Knowledge of the composition and structure of the biological communities of these wetlands is only one component that would help to determine the conservation status of these ecosystems (Figueroa et al., 2009). For this reason, conservation actions should focus on the synergistic effects of climate change and its interaction with the multiple threats to biodiversity, including habitat degradation, overexploitation, and invasive species (Brook et al., 2008).

3.4 Conservation actions in the Aguada La Chimba urban coastal wetland

A conservation project is currently underway in the Aguada La Chimba wetland. This project started in 2018 with a baseline assessment and the identification of threats, and was later consolidated with the declaration of the wetland as a Natural Sanctuary (DS No. 14/2021 MMA) and urban wetland (Res. No. 2). These legal figures of protection are warranted given the value of coastal waters as a refuge for local biodiversity and migratory birds. Among the biological conservation elements identified by the SNASPE (Servicio Nacional de Áreas Silvestres Protegidas del Estado [National System of Wildlife Areas Protected by the State of Chile]) are the snail from La Chimba (Heleobia chimbaensis), an endemic species categorized as Vulnerable (RCE, DS 52 MMA, 2014); Chimbanito stream; El Rubio stream, and the migratory bird assemblage. As for the latter, some migratory birds have been reported in the wetland, such as sanderlings (Calidris alba), which have a varied diet that includes arthropods and polychaetes (Petracci, 2002; Castro et al., 2009; Grond et al., 2015). Since wetlands concentrate a large number of birds that inhabit these areas depending on food availability, it is important to better understand this group.

Some of the major threats to these biological conservation elements include the illegal extraction of water and the intervention of water courses; the presence of exotic invasive species (Canis lupus familiaris, Oryctolagus cuniculus, Melanoide tuberculata, and Cavia porcellus) that displace or predate on native species; habitat fragmentation, and the illegal dumping of waste and rubble. The control, mitigation and eradication of these threats will be incorporated into a management plan intended to recover the wetland’s ecological attributes with the goal of restoring the ecosystemic services that are necessary to preserve the local biodiversity in the long run. Embedded in an urban matrix, the Aguada de La Chimba wetland is a special case since it is interacts with the fauna associated to these artificial systems (rabbits, rats, and dogs). However, since this wetland still hosts native and endemic species, recovering it and controlling its threats may set a precedent in terms of natural heritage conservation and may turn it into a hotspot and refuge for local biodiversity.

One of the major current challenges is coordinating the different territory stakeholders involved in monitoring and controlling the threats and activities that are incompatible with the goal of preserving the wetland area. Also, the lack of resources for conducting conservation actions makes it necessary to prioritize the most urgent ones. And although the governance model based on public–private partnerships and the engagement of the community has been successful, agreements are still required to ensure these relationships persist in time and are made operable.

Apart from the in-situ conservation actions intended to control threats in coastal wetlands, the enforcement of current regulatory instruments is required, including the definition of soil uses that are compatible with environmental protection efforts, to support the measures takes by government agencies and territory stakeholders. This must also be supported by profound regulatory changes aimed at preventing damage to wetlands, such as the prohibition of extracting surface and underground waters, an action which has a direct impact on the ecosystem and the arthropod habitats.