Non-native herpetofauna continue to proliferate in the world’s most invaded herpetofauna community: Evidence against community saturation

Non-native herpetofauna continue to proliferate in the world’s most invaded herpetofauna community: Evidence against community saturation | Authorea try { document.documentElement.classList.add('js'); } catch (e) { } var _gaq = _gaq || []; _gaq.push(['_setAccount', 'G-8VDV14Y67G']); _gaq.push(['_trackPageview']); (function() { var ga = document.createElement('script'); ga.type = 'text/javascript'; ga.async = true; ga.src = ('https:' == document.location.protocol ? 'https://ssl' : 'http://www') + '.google-analytics.com/ga.js'; var s = document.getElementsByTagName('script')[0]; s.parentNode.insertBefore(ga, s); })(); Skip to main content Preprints Collections Wiley Open Research IET Open Research Ecological Society of Japan All Collections About About Authorea FAQs Contact Us Quick Search anywhere Search for preprint articles, keywords, etc. Search Search ADVANCED SEARCH SCROLL Ecology and Evolution This is a preprint and has not been peer reviewed. Data may be preliminary. 17 January 2025 V1 Latest version Share on Non-native herpetofauna continue to proliferate in the world’s most invaded herpetofauna community: Evidence against community saturation Authors : Stephanie L. Clements 0000-0002-3989-6030 , F. Michael Ackerman 0000-0003-1255-3123 , Isabella M. Olensky , Elizabeth White , Millie E. Rogers , and Christopher A Searcy 0009-0004-1026-4015 [email protected] Authors Info & Affiliations https://doi.org/10.22541/au.173710805.55043337/v1 476 views 389 downloads Contents Abstract Information & Authors Metrics & Citations View Options References Figures Tables Media Share Abstract The spread of non-native species continues to increase around the globe, making it important we understand the dynamics of the resulting communities in which non-natives comprise a high percentage of the total fauna. As non-native species continue to invade, the resulting community may become saturated, at which point limited resources would prevent colonization by new non-native species and native or already established species might decline. As the global hotspot for non-native reptiles and amphibians, South Florida’s herpetofaunal community has a higher probability of having reached the saturation point than any other comparable system. Surveys conducted in Miami-Dade County in 2017 demonstrated that non-native species already dominated both native and non-native habitat types and provided a baseline to examine dynamic changes such as signatures of community saturation or negative impacts on native species. In 2022, we replicated the surveys from 2017 at the same 30 sites. We found that non-native richness and abundance have increased significantly (19% and 33% increase in overall alpha diversity and abundance, respectively), showing no signs of community saturation. We also found no correlation between these non-native increases and decreases in either native species richness or abundance. Non-native species richness increased more rapidly at sites dominated by non-native habitats, with two rock-loving species, Agama picticauda and Leiocephalus carinatus, standing out as the most rapidly spreading non-native herpetofauna. Our findings demonstrate that open niche space allows the continued expansion of non-native herpetofaunal populations even in the highly invaded community of Miami-Dade County, and that protection of native habitat may help slow the spread of non-native species. Introduction Human population growth, increased urbanization, and other anthropogenic activities have led to the spread of non-native species on a global scale (Perrings et al. 2005; Marques et al. 2020; Banks et al. 2015; Jeschke et al. 2014; Pyšek et al. 2020). This worsening spread and increased abundance of non-native species (Pyšek et al. 2020) makes it critical that we understand how non-native species are changing ecological communities and impacting native species (Seebens et al. 2017). As non-native species move into new communities, these communities could reach a saturation point at which either these non-native species cannot continue to encroach, or native species are lost to make room for new non-natives (Starzomski et al. 2008; Pinto-Sanchez et al. 2014; Sax & Gaines 2008). The concept of community saturation stems from competition for niche space (Starzomski et al. 2008; Macarthur & Levins 1967; Vanni et al. 2009). Ecologists have long known that there is an upper limit to the number of species that can exist in a system with finite resources; however, it can be difficult to determine whether that limit has been reached (Olivares et al. 2018). Within invaded communities, saturation can occur in two ways in terms of species richness: extinction-based or colonization-based saturation (Sax & Gaines 2008). At the point of saturation in an extinction-based framework, the addition of a new non-native species will result in the local extinction of an already established species (Sax & Gaines 2008; MacArthur & Wilson 1967). Alternatively, in a colonization-based framework, at the point of saturation the addition of a new non-native species is inhibited by established species due to a limitation of niche space (Sax & Gaines 2008; Tilman 2004). Therefore, when community saturation is occurring, either non-native species replace native species (Pinto-Sanchez et al. 2014) or non-native species are prevented from establishing. If community saturation is not occurring, there is room for community expansion and non-native species can proliferate without replacing already established populations. Previous studies have seen mixed results on whether communities are saturated. Some invaded systems have demonstrated that non-native species can continue to increase without decreasing native richness (Thomas & Palmer 2015; Davis et al. 2015), while others have demonstrated a loss of native species as non-native species encroach (Honek et al. 2016; Haubrock et al. 2020). Yet other studies have demonstrated native species outcompeting non-native species, preventing establishment by the non-native species (Tuckett et al. 2021; Thompson et al. 2012). Whether or not a community reaches a point of saturation, non-native species may have negative impacts on native populations. For example, within numerous systems, non-native species have been documented as one of the chief reasons for declines in native species richness (Dorcas et al. 2012), shifts in native species composition (Honek et al. 2016), and community homogenization (Bando et al. 2023). Additionally, hybridization between native and non-native individuals can result in the replacement of native populations with hybridized individuals (Huxel 1999; Kraus 2015). Negative ecological impacts of non-native species include predation on native species, the transmission of diseases, and increased resource competition (Rodda & Savidge 2007; Picco & Collins 2008; Cole & Harris 2011; Kraus 2015). The last of these is most likely to occur between species with similar ecological roles and, given phylogenetic signal in such traits, is most often observed between native/non-native congeneric pairs (Gioria & Osborne 2014; Zwerschke et al. 2018). Considered a global hotspot for invasive and non-native species, Florida has the largest number of established non-native amphibian and reptile species in the world (Capinha et al. 2017), many of which have been present in the state for several decades. Most non-native species occur in South Florida. The oldest introductions occurred ~150 years ago with the brown anole ( Anolis sagrei ) introduced in 1887 (Garman 1887) and the greenhouse frog ( Eleutherodactylus planirostri s) introduced in 1863 (Cope 1863), leaving ample time (~136 generations for brown anoles, ~240 generations for greenhouse frogs; Lee et al. 1989; Meshaka & Layne 2002) for population expansion. As such, South Florida’s highly invaded herpetofaunal community has a higher probability of showing saturation or negative effects on native species than other similar systems around the globe. In 1958, only 12 non-native herpetofauna species were recorded in Florida compared to the 63 recorded as of 2016, highlighting the continuing dramatic increases in non-natives in this region (Duellman & Schwartz, 1958; Krysko et al. 2016). Numerous factors contribute to the large number of non-native species in Florida, including the subtropical climate, high levels of natural disturbance, and the thriving exotic species trade (Smith, 2005; Krysko, 2009; Fujisaki et al., 2009; Searcy et al. 2023). Further, the loss of native habitat in South Florida due to land-use change continually threatens native biodiversity and provides a foundation for non-native species to flourish (Divya et al. 2021; Marques et al. 2020). Despite initiatives to increase the footprint of protected native habitat in Miami (Diamond & Heinen 2016; Alonso & Heinen 2011), urban development has increased 200% over the last 20 years (Divya et al. 2021). All these factors contribute to South Florida being the most highly invaded continental ecoregion in the world (Searcy et al. 2023). There are substantial records of non-native species introductions in Florida, as well as of increases in the abundance of non-native species (Krysko et al. 2009; Krysko et al. 2011; Krysko et al. 2016). However, there are limited studies that have investigated changes in local abundance and alpha diversity, where one is more likely to observe community saturation than at a regional level. Cassani et al. (2015) scratches the surface of investigating change over time by finding significant increases in the brown anole, Anolis sagrei , and concluding that herpetofaunal communities are changing rapidly in their study region of southwest Florida. However, this study was conducted in a large native habitat parcel and not in urbanized areas (Cassani et al. 2015). We are more likely to detect saturation and negative effects on natives in urbanized systems, such as fragmented habitat patches within urbanized Miami-Dade County, where there is a greater prevalence of non-native species (Byers 2002; Clements et al. 2019; Marques et al. 2020). A survey completed in 2017 revealed that non-native species dominated the herpetofaunal community in both non-native and native habitats in Miami-Dade County (Clements et al. 2019). This previous survey, spanning 30 sites, provides a baseline to allow for investigation of how the herpetofaunal community is changing over time, whether it is showing signs of saturation at the local scale, and whether non-native species are having an impact on native species richness or abundance. To address these questions, we replicated the surveys from 2017 at the same sites and during the same time of year (Clements et al. 2019). The ability to re-survey the same sites presents us with a unique opportunity to investigate the changes in the herpetofaunal community of South Florida, the global hotspot for non-native species, over a span of just five years. Site selection Survey sites replicated those from Clements et al. (2019) and included 15 urban green spaces (non-native parks) and 15 parks where the native habitat has been preserved with >50% native vegetation (native parks) across Miami-Dade County. Native and non-native parks were paired geographically and spanned the same range of fragment sizes from 0.6 to 229 ha. For a detailed description of site selection see Clements et al. (2019). Survey methods We conducted diurnal visual encounter surveys from March to May 2022 (peak local reptile and amphibian activity season), replicating the active survey methods performed five years earlier in 2017 (Clements et al. 2019). Visual encounter surveys (Enge et al. 2004b) were performed between 8 am and 7 pm during favorable weather conditions (e.g., minimal rain and temperatures above 18°C). This is the same set of weather conditions used to select 2017 survey dates, and thus while there was some variation in weather conditions between surveys within each year, mean conditions in 2017 and 2022 should be comparable. The time spent surveying each park replicated that of the 2017 surveys such that larger parks were sampled to the same depth on the species abundance curve as smaller parks (Clements et al. 2019). While each park was only surveyed once, this required that some of the larger parks were visited on multiple days to cover the entire area. Some parks had both native and non-native habitats, but only native habitats were surveyed in parks with majority native habitat and only non-native habitats were surveyed in parks with majority non-native habitat, although we did survey transition zones along habitat edges. These choices mimicked those made in the 2017 surveys (Clements et al. 2019) because the goal of the 2022 surveys was to detect changes in the herpetofaunal community over the intervening five-year period. The habitats within the parks did not notably change over the course of the 5-year period. Altogether, the surveys involved 384 person hours conducted by a team of 13 surveyors trained in identification of local reptile and amphibian species. Survey hours were not divided evenly by the 13 surveyors (minimum 3 hours, maximum 114 hours) as they were dependent on availability. More than 80% of all survey hours were completed by those listed as authors on this manuscript, and the lead author was present at all sites to ensure consistent survey techniques were followed both within the 2022 surveys and with regard to the 2017 surveys. We found animals by sound and sight, scanning trees, tall grasses, bushes, anthropogenic structures within the parks, and other perching surfaces. Water bodies were surveyed visually, with particular attention to the shoreline. When necessary, we used binoculars to identify individuals out in water bodies or up in trees. We also searched under cover objects such as palm fronds, rocks, and anthropogenic materials (e.g., trash cans, etc.). When multiple habitat types (e.g., mangroves, pine rockland) were present in a park, all habitat types were thoroughly examined, within the relevant native/non-native classification of the park. Animals were only captured if necessary for identification and were then immediately released. Handling of animals was approved under University of Miami IACUC protocol 22-016 in accordance with the ASIH/HL/SSAR Guidelines for use of live amphibians and reptiles in field research. Statistical analyses As in Clements et al. (2019), we used a second-order jackknife to calculate estimated species richness in each park and the percentage of the estimated species that we observed (Gotelli and Colwell 2011). To ensure that our sampling in 2022 was to the same depth on the species abundance curve as in the 2017 surveys, we used a paired t-test to determine if there was a difference in the percentage of estimated species that we observed between years, which could indicate a change in detectability. We then investigated whether there was a difference in herpetofaunal abundance or richness based on survey year or habitat type (native/non-native) or their interaction. To address this question, we conducted factorial ANOVAs with year, habitat type, and their interaction as fixed effects and with park as a random effect to compare: (1) overall abundance and richness, (2) native abundance and richness, (3) non-native abundance and richness, and (4) proportion of native individuals. Abundances and native richness were log-transformed and the proportion of native individuals was arcsine-transformed for normality. These analyses were conducted in JMP 15 (JMP® Pro, Version 15.0.0, 2019). To determine whether other environmental factors influenced the richness or abundance of non-native herpetofauna in 2022, we conducted an ANCOVA with 2017 non-native richness (or abundance) as a covariate. Environmental factors that we considered included park area, native species richness (or abundance), habitat type, and connectivity. We evaluated all possible models and selected the model with the lowest AICc using JMP 15. Connectivity was obtained from Clements et al. (2019), and was calculated using a dispersal constant of 0.1. Area, connectivity, and abundances were log-transformed for normality. No models showed evidence of spatial autocorrelation ( P > 0.3), which was tested in R Version 4.0.2 using the lctools package with the number of nearest neighbors set to five (Ghosh 2006; Kalogirou 2017; R Core Team 2022). We also investigated whether the temporal change (2017 vs. 2022) in either abundance or richness was correlated between native and non-native species using a Pearson correlation (data were normally distributed). This should illuminate whether there is a net negative effect of non-native species accumulation on native herpetofauna. We followed this by looking at some individual species interactions between close relatives or common species. All of the correlations between temporal changes in individual species abundances were tested using Spearman’s correlations for non-parametric data. We looked at the following species pairs: 1) Anolis sagrei and A. carolinensis , the most common non-native and native species, respectively, 2) A. carolinensis and A. cristatellus , 3) A. sagrei and A. cristatellus , 4) A. sagrei and Agama picticauda , 5) A. sagrei and Leiocephalus carinatus , 6) A. carolinensis and A. picticauda , 7) A. carolinensis and L. carinatus , and 8) A. picticauda and L. carinatus . The first three comparisons look at competition between native/non-native pairs of arboreal lizards with similar niches (Edwards and Lailvaux 2012), the next four comparisons look at the most common native and non-native lizards and the most rapidly spreading predatory lizard species (Schoener et al. 2017), and the last comparison is between these rapidly spreading predators. All these analyses were conducted in JMP 15. We wanted to ascertain if there was a difference in community composition between native and non-native habitats, as there was in Clements et al. (2019), and whether the indicator species for native and non-native habitats remained the same as they were five years ago. We conducted a PERMANOVA using the Bray-Curtis distance metric (Anderson 2001; Clements et al. 2019) to determine if there was a difference in community composition between native and non-native parks using the new 2022 dataset. Any species found at a single park were excluded from the analysis, except for species that could be combined into a monophyletic group, such as Hemidactylus spp. (Clements et al. 2019). We relativized the community matrix to be consistent with the PERMANOVA analysis in Clements et al. (2019) by relativizing each row by person-hours of survey time and each column by its total abundance to upweight rare species. We followed this PERMANOVA with an indicator species analysis to determine which species were associated with native versus non-native parks (De Caceres and Legendre 2009). We next determined whether there was a difference in community composition between survey years, again using a PERMANOVA. For this PERMANOVA, we combined our 2017 and 2022 data sets, and then dropped any species that were found in only one park*year combination and could not be combined into a monophyletic group (i.e, Hemidactylus spp., Nerodia spp., Plestiodon spp., and Pseudemys spp.). Similarly, we corrected taxonomic names that differed between years to correct species and collapse columns. For example, we had labeled blindsnakes “ Rhamphotyphlops braminus ” in 2017, but the currently accepted nomenclature is “ Indotyphlops braminus ”. We also misidentified A. picticauda as Agama agama in 2017, so we corrected this column from the 2017 data. We did not relativize this combined matrix by row or column because we were interested in comparing the two survey years and had used consistent survey effort to collect the data in both years. With this combined matrix, we ran a PERMANOVA with Bray-Curtis distance metric and blocking by park. We followed this with an indicator species analysis to determine which species were most associated with 2017 versus 2022. All PERMANOVA and indicator species analyses were conducted in R Version 4.0.2. Lastly, we were interested in whether we could identify predictors of abundance change for the two species that have proliferated most rapidly over the last five years: A. picticauda and L. carinatus . We used the same ANCOVA model selection described above to evaluate whether habitat type, connectivity, or park area helped to predict the 2022 abundance of these species, with the covariate of their abundance in 2017. Abundances, connectivity, and area were log-transformed, and analysis was conducted in JMP 15. We also evaluated whether there was a spatial pattern to the increase in these species by looking for spatial autocorrelation, which was tested in R Version 4.0.2 using the lctools package with the number of nearest neighbors set to five (Ghosh 2006; Kalogirou 2017). Results In total we recorded 9,535 individuals across 36 species in 2022. This represents an increase in individuals of 30% and in species of 14% over Clements et. al (2019). Only 7.7% of the individuals in the 2022 dataset were native, compared to 9.4% in Clements et. al (2019). Of the species recorded, 47% were native and 53% were non-native. Based on the second-order jackknife, there was no difference in the proportion of species observed between the 2017 and 2022 surveys (P = 0.14). Both overall abundance (P = 0.036) and richness (P = 0.0006) of herpetofauna increased significantly in the five years between surveys. However, when separating native and non-native species, we see this increase is only true for non-native species, which increased 19% in richness (P < 0.0001) and 33% in abundance (P = 0.027). Native species richness (P = 0.68) and abundance (P = 0.34) did not notably differ from the 2017 surveys (figure 1). There was a marginally significant interaction between habitat type and year on non-native richness (P = 0.054) in the direction of non-native species increasing more rapidly in non-native parks, but otherwise habitat type did not have a significant impact on richness, abundance, or proportion of native individuals (P > 0.09). However, ANCOVA revealed that 2022 non-native richness was higher in non-native parks (estimate = -0.84, P = 0.009; figure 2a), and in parks with greater 2022 native richness (estimate = 0.59, P = 0.03), as well as where there was a higher non-native richness in 2017 (estimate = 0.59, P = 0.002) (Best model AICc = 118, P < 0.0001, R 2 = 0.68). Non-native abundance in 2022 was best predicted by 2017 non-native abundance (estimate = 0.44, P = 0.0073) and 2022 native abundance (estimate = 0.49, P = 0.0019), but not habitat type (Best model AICc = 50, P < 0.0001, R 2 = 0.80; figure 2b). Park size and connectivity were not selected as important predictors for non-native richness or abundance. Species analyses As was true in 2017, the majority of the individuals recorded were from the genus Anolis (82% in 2022 and 86% in 2017). Non-native species from genus Anolis accounted for 75% of total individuals recorded. The most abundant herpetofauna species were Anolis sagrei (43.8%), Anolis cristatellus (18.0%), and Anolis distichus (12.3%). The other abundant non-native species recorded were Agama picticauda (3.8%), Rhinella marina (3.1%), Iguana iguana (2.4%), Ctenosaura similis (2.4%), Basiliscus vittatus (1.9%), and Leicephalus carinatus (0.9%). The most abundant native species were Anolis carolinensis (6.7%), Sphaerodactylus notatus (0.4%) and Coluber constrictor (0.3%). The most widespread native species were A. carolinensis (97% of parks surveyed) and C. constrictor (43%), and the most widespread non-native species were A. sagrei (97%), A. distichus (87%), Hemidactylus spp. (70%), A. picticauda (57%), A. equestris (53%), Indotyphlops braminus (50%), B. vittatus (47%), I. iguana (43%), A. cristatellus (40%), Eleutherodactylus planirostris (33%), Leiocephalus carinatus (27%), Osteopilus septentrionalis (17%), and Trachemys scripta (17%). Five species, A. cristatellus , A. picticauda, E. planirostris , I. iguana , and L. carinatus were found in ≥5 more parks in 2022 than in 2017. Two species, Pseudemys nelsoni and Rhinella marina , were found in ≥5 fewer parks in 2022 than they were in 2017. Of the 15 species found in at least five parks, only two were native. There was no correlation between the change in abundance of native and non-native species (P = 0.31, R 2 = 0.04), nor between the change in richness of native and non-native species (P = 0.65; R 2 = 0.008), indicating that the increase in richness and abundance of non-natives is not having a detectable impact on the richness or abundance of natives (figure 3). When we broke this down by species, there was also no correlation between the change in abundance for any of the species interactions tested (P > 0.10), with the exception of A. picticauda and L. carinatus , which increased together (P = 0.016; Spearman’s ρ = 0.43). Our 2022 PERMANOVA revealed a significant difference in community composition between native and non-native habitats (P = 0.001), as was also true in the 2017 surveys (P = 0.001; Clements et al. 2019). However, our indicator species differed from those found in the prior survey (Clements et al. 2019). Our 2022 data revealed that a native snake, C. constrictor , was the indicator species for native habitats (P = 0.03), whereas in 2017, the indicator species of native habitat was the non-native snake I. braminus (Clements et al. 2019). In 2022, we also found four indicator species for non-native habitat: A. equestris (P = 0.0001), A. picticuada (P = 0.0002) , A. distichus (P = 0.002) , and L. carinatus (P = 0.006) . In 2017, A. sagrei and A. equestris were indicator species of non-native habitat (Clements et al. 2019). A separate PERMANOVA found a significant change in community composition through time (P = 0.002; figure 4). The indicator species for 2022 were A. picticauda (P = 0.0001) and L. carinatus (P = 0.028), while there were no indicator species for 2017. Because A. picticauda and L. carinatus were revealed as the indicator species for 2022, and also were indicator species of non-native parks in 2022, but not 2017, we wanted to further investigate the increase in these two species. While there was an average of just 0.33 A. picticauda per park in 2017 (Clements et al. 2019), we found an average of 11.9 A. picticauda per park in 2022 (a 36-fold increase). Even more notably, A. picticauda were found in only 2 parks in 2017 and in 17 parks in 2022. In 2017, L. carinatus was found in just 2 parks and an average of 0.13 individuals per park, whereas in 2022 we found L. carinatus in 8 parks and an average of 2.90 L. carinatus per park (a 22-fold increase). When blocking by park, the abundance of A. picticauda was significantly higher in 2022 (P = 0.02), while the abundance of L. carinatus was only marginally higher in 2022 (P = 0.07). Through model selection, we found that the 2022 A. picticauda abundance was higher in non-native parks (estimate = -0.7, P = 0.004) and parks with a higher abundance of A. picticauda in 2017 (estimate = 1.05, P = 0.06; best model AICc = 103, P = 0.001). The 2022 L. carinatus abundance was also higher in non-native parks (estimate = -0.32, P = 0.04) and parks with a higher abundance of L. carinatus in 2017 (estimate = 1.95, P = 0.001; best model AICc = 76.8, P = 0.0002). There was no detectable geographic pattern or spatial autocorrelation between parks with the greatest increase in A. picticauda or L. carinatus (P > 0.20). However, anecdotally, the authors believe that there is a relationship between parks where agamas are increasing and presence of cement or other hard man-made structures either inside or along the edge of the park, which aligns with recent findings (Mothes and Searcy 2024 ). Discussion Our surveys indicate that the non-native herpetofaunal community of Miami-Dade County has experienced rapid increases in both abundance and richness within just five years. Between 2017 and 2022, we found that the richness of non-native species increased across both native and non-native parks by an average of 1.4 species per park. This appears to be a true increase in richness rather than an increase in detectability, as the percentage of the predicted species pool observed did not change significantly between years. In addition, because the increase in actual species observed was attributed solely to the non-natives, we believe this represents a real expansion in the non-native community and is a considerable increase in non-native richness in a relatively short period of time. We also found that the abundance of non-native species increased across both native and non-native parks. While there are certainly differences in detectability across herpetofauna species based on behavior and activity, those differences should be consistent between 2017 and 2022, making our abundance comparisons across time valid. Compared to the 2017 surveys, non-native abundance increased by 32.7% (a significant increase) while native abundance increased by only 6.5% (a non-significant change). While it is feasible that the increased abundances detected in 2022 were a result of the surveyors being more skilled, there is no obvious reason why this would be true (the surveyors were mostly graduate students in 2017 and undergraduates in 2022), and if this was the driving factor we would expect increases in natives and non-natives to be of similar magnitude. In the 2022 surveys, 92.3% of individuals were non-native (compared to 90.6% in 2017; Clements et al. 2019). Even among the 7.7% of individuals that we recorded as native, there is an important caveat to consider. The most abundant native species recorded was the green anole ( A. carolinensis ), but studies have reported that the green anole population in Miami-Dade County consists of morphologically similar hybrids between the native A. carolinensis and the non-native Cuban green anole ( Anolis porcatus ; Wegener et al. 2019). If we were to reclassify A. carolinensis as non-native due to this hybridization, then non-natives would make up an astounding 99% of all individuals recorded. This staggering abundance of non-native individuals shows why South Florida provides unique insights into the spread and effects of non-native species (Fujisake et al. 2009). Over the last five years, non-native species richness in this ecosystem has increased by 18% while non-native abundance has increased by an even greater 33%. These results indicate that the South Florida herpetofaunal community has not yet reached a saturation point, as non-natives continue to increase in both local abundance and how widespread they are. This trend is also seen in fish assemblages across North American freshwater ecosystems and in plant assemblages in the Pacific Northwest, which have similarly experienced rapidly increasing non-native species richness yet show no evidence of reaching regional community saturation (Stohlgren et al. 2008; Mitchell et al. 2009). On a global scale, the increase in distribution of non-native species also does not show any sign of saturation (Seebens et al. 2017). South Florida represents an ecosystem where the long history and remarkable abundance of non-natives makes it more likely than most to approach saturation (Searcy et al. 2023). It is therefore notable that, even in this global hotspot of introduced biodiversity, there is no sign of community saturation even at the scale of individual sites (i.e., the scale at which species interactions occur). With both non-native abundance and richness on the rise, we expected that we might see negative impacts on native species. However, at least at the community level, there is no clear signature of such an effect. Native abundance and richness remained relatively constant over the last five years at each of our study sites, regardless of increases in non-native abundance and richness. While lack of a non-native effect on native species has certainly been documented in other systems (e.g., herbaceous plants; Davis et al. 2015), it is counter to the global trend. In general, native responses to invasion become more negative as both the abundance and trophic level of the invading species increases (Gleditsch et al. 2010; Bradley et al. 2019). It is therefore interesting that, despite a high abundance of invading species spanning many trophic levels, there is not a negative impact of invaders on the richness or abundance of native herpetofauna in South Florida. It is important to note that measurements of richness and abundance look at detrimental effects on native species at the community level, but that individual non-native species may still be affecting individual native species even where community level responses are not seen. For example, the invasion of cane toads ( Rhinella marina ) in northern Australia resulted in population declines in several species of native predatory lizards ( Varanus panoptes , V. mertensi , and V. mitchelli ), but also a population increase of one of V. panoptes ’s native prey species, the lizard Amphibolurus gilberti (Doody et al. 2009). In this case, overall native abundance would not have reflected the severity of native predator species’ declines due to the offsetting increase in the native prey species. We tested for non-native effects at the individual species level, but we found no correlation between changes in abundance of any two individual species (other than a positive correlation between A. picticauda and L. carinatus , both non-native species) . However, many native species were only detected in these surveys at very low abundances or at very few locations, limiting replicability across sites and therefore our ability to detect significant correlations. We did investigate whether there was a correlation between the change in the most abundant non-native, Anolis sagrei , and the most abundant native, A. carolinensis , but found no evidence of this. However, we only looked at changes in abundance, while other studies have demonstrated that A. carolinensis may be affected behaviorally by non-native encroachment. The native species demonstrates a niche shift toward greater arboreality in the presence of A. sagrei (Edwards and Lailvaux 2012; Stuart et al. 2014). This behavioral and associated morphological (Glossip and Losos 1997; Stuart et al. 2014) shift demonstrates adaptation by a native species to avoid direct negative effects from the encroachment of a non-native. Some non-natives are notorious for their destructive effects on native species, such as Python bivittatus and Salvator merianae. We detected both species at our survey sites for the first time in 2022, despite ongoing conservation efforts to contain and eradicate populations of these species. Significant declines in native mammals have been documented in Everglades National Park, particularly in areas with high P. bivittatus proliferation (Dorcas et al. 2012). Like P. bivittatus , S. merinae is a generalist predator, preying on American alligator ( Alligator mississippiensis ) and red-bellied cooter ( Pseudemys nelsoni ) eggs (Mazzotti et al. 2015). In 2021, the Florida Fish and Wildlife Conservation Commission (FWC) introduced new regulations regarding the keeping, breeding, trading, and selling of 16 prohibited reptile species, including P. bivittatus and S. merinae (FLA 2021). However, these measures have not stopped the dispersal of individuals from well-established populations. Furthermore, non-native species can have more subtle impacts on native species, such as the spillover of invasive parasites (e.g., spillover of a harmful invasive Asian parasite ( R. orientalis ) from Burmese pythons to 14 native snake species; Miller et al. 2017; Miller et al. 2020). While beyond the scope of our study, it is important to recognize that non-native impacts on native species extend beyond decreases in abundance or richness. While we are most concerned with the impact that non-native species may have on native species, we also investigated potential correlations between changes in non-native species, as the loss of a non-native species at a site could be a sign of community saturation. This was previously evidenced by ecologically analogous non-native Hemidactylus spp.’s inability to stably co-exist, resulting in the competitively dominant H. mabouia rapidly displacing H. garnotii in their sympatric range in central and southern Florida (Meshaka 2000; Short and Petren 2012). However, we did not see any evidence in our surveys of a negative correlation between the non-native species pairs that we investigated. Again, this may be due to few co-occurrences of certain species preventing the replicability needed to test this effect. For example, at Evelyn Greer Park, the number of Ameiva ameiva observed decreased dramatically from 2017 to 2022, while A. picticauda were detected there for the first time in 2022, and at a high abundance. While this could indicate a potential displacement of A. ameiva by A. picticauda , the two species do not co-occur at enough sites to test this hypothesis. Similarly, on multiple occasions, we observed A. picticauda preying on non-native Anolis species, and pursuing the native A. carolinensis . As a large predatory lizard, A. picticauda could negatively impact populations of smaller species, although we could not detect such an effect. We also investigated the impact of habitat type on non-native increases in richness and abundance. Clements et al. (2019) found no significant difference in native or non-native herpetofaunal abundance or richness based on habitat type. While it remains true that there is no difference in richness or abundance of native species between habitat types in 2022, we now see that the increase in non-native species is more rapid in non-native habitats. Disturbances to native habitats have been shown to facilitate the diversity and abundance of non-native species in plant systems as well (Jauni et al. 2015). Given that there was no difference in non-native penetration of native vs. non-native habitat during the 2017 surveys, there must have been some previous time point in the herpetofauna invasion of South Florida when non-native species were not inhibited by native habitat. However, for the species currently expanding in Miami, it seems to be the case that the native parks serve as an impediment to spread, demonstrating an important benefit of the preservation of native parks, which may also serve as a reservoir for native species, similar to what has been seen in other systems (Chace et al. 2006). In our study, we saw that the indicator species of native parks in 2022 was the native Coluber constrictor , demonstrating that these native habitat areas may be a refuge for common native species. The primary indicator species for the change in community composition from 2017 to 2022 were Agama picticauda and Leiocephalus carinatus , non-native species that have increased in abundance 36-fold and 22-fold, respectively, since 2017. These two species were also new indicator species for non-native parks in 2022. Taken together, these results demonstrate that these two non-native species have increased rapidly in distribution and abundance in non-native habitat areas across Miami-Dade County in just the last 5 years. The current A. picticauda population is thought to have been introduced in the Homestead area in 1992 as a result of Hurricane Andrew (Enge et al. 2004), with DNA analyses indicating multiple subsequent introductions contributing to their current genetic diversity (Nuñez et al. 2016). Leiocephalus carinatus was first reported in Florida in Palm Beach County in 1958 and has subsequently spread towards Miami (Smith et al. 2004). Both L. carinatus and A. picticauda are predisposed to succeed in human-disturbed environments, as in their native range both species prefer open habitats with rocky structures for basking (Neel et al. 2020; James et al 1979). In urban environments, impervious surfaces, such as parking lots, provide a suitable substitute for their preferred habitats (Meshaka et al. 2022). The abundance of A. picticuada can also be positively predicted by the presence of human structures such as dumpsters (Mitchell et al. 2021), which likely serve as refugia as well as attracting prey. Non-native parks contain plentiful warm stone basking locations including curbs, parking lots, and sidewalks, while also providing shade and protection from predators through shrubbery, trees, etc. (Moore et al. 2006). Both A. picticauda and L. carinatus are clearly dispersing rapidly, as A. picticauda and L. carinatus were both detected in only 6.6% of parks in 2017 yet were found in 57% and 27% of parks, respectively, in 2022. These species likely use similar dispersal methods to other local non-native lizard species, especially the extremely widespread and abundant A. sagrei, for which vehicular rafting is a theorized dispersal method (Campbell 1996). Similarly, A. picticauda has also been documented using cars and interstates for dispersal (Moore 2019) and is speculated to hitchhike on freight using railways (Gray 2020). There is also evidence for dispersal of both L. carinatus and A. picticauda being aided by the nursery industry, as researchers have witnessed L. carinatus perched on landscaping vegetation piles (Smith et al. 2004) and 72% of the suspected origins of A. picticauda subpopulations are located within 0.5 km of plant nurseries, importers, or exporters (Gray 2020). These modes of dispersal help explain the increase in non-native species in non-native habitats, as native habitats experience less landscaping and park-goers, as well as fewer other hallmarks of urbanization that would facilitate the spread of these species. It is worth noting, however, that all native parks in this study are within the urban matrix and even for the largest ones none of their interiors are more than 700 m from an urban area (Clements et al. 2019), which may play an important role in why we see such a high abundance of non-natives even in our native park sites. Despite a long history of non-native herpetofauna establishment in South Florida, non-native richness and abundance have continued to increase over the last five years, while native richness and abundance has remained relatively constant. As such, there does not appear to be a negative impact of non-natives on native richness or abundance over the five-year time frame in which this study was conducted. Further studies will be necessary to determine if this remains true or if native populations begin to decline more widely because of direct competition/predation from non-natives beyond the losses that have already occurred due to loss of habitat. It is possible that the native populations experienced a significant decrease at some point in the past, such as during the initial rise of many of these non-natives. We know that non-natives can cause rapid decreases in native populations and even extinctions in a less than five-year time period, such as was the case with the non-native brown tree snake in Guam (Savidge 1987; Rodda et al. 1997). As discussed in Clements et al. (2019), there are some native species in Miami-Dade County, such as skinks ( Plestiodon spp.), that are now recorded at lower abundances and/or occurrences than surveys from the early 2000s seem to suggest (Enge et al. 2004). Our study would not be able to determine if native species were already lost, or had reached a new, but lower, stable population size prior to the 2017 surveys. Similarly, our study cannot assess whether there has been any change in native or non-native abundance/richness in larger, contiguous native habitat areas since our focus was on urban Miami. Future studies should investigate these dynamics in larger natural areas such as the Everglades, where the abundance of non-native species remains lower than in the urban core (~16% non-natives in Loxahatchee National Wildlife Refuge; Howell et al. 2021). Native abundance and richness did not differ significantly between non-native and native parks over the five years, but non-native parks demonstrated the highest increase in non-native richness. This suggests that both habitats are important for herpetofauna and for contributing to native diversity, but prioritizing the conservation of native habitats will be critical when attempting to inhibit the spread of non-native species. 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Conflict of Interest : The authors declare that they have no conflict of interest. Data Availability Statement: The data that support the findings of this study are available in the supplementary material of this article. Research Involving Animals : Procedures involving animals were approved under IACUC protocol number 22-016 at the University of Miami in accordance with the ASIH/HL/SSAR Guidelines for use of live amphibians and reptiles in field research. Permits : This work was conducted under a permit from Florida Fish and Wildlife Conservation Commission (#LSSC-16-0013C) and under a permit from Miami-Dade County Parks and Recreation Department (#345). Author Contributions: Project conception and design were led by SLC and CAS. All authors contributed to the data collection. Analysis and data management were performed by SLC and FMA. Figures were designed by SLC. Funding was secured by CAS. The first draft of the manuscript was written by all authors collaboratively. All authors commented on previous versions of the manuscript and all authors read and approved the final manuscript. Information & Authors Information Version history V1 Version 1 17 January 2025 Copyright This work is licensed under a Non Exclusive No Reuse License. Collection Ecology and Evolution Keywords community ecology statistical terrestrial vertebrate Authors Affiliations Stephanie L. Clements 0000-0002-3989-6030 University of Miami College of Arts and Sciences View all articles by this author F. Michael Ackerman 0000-0003-1255-3123 University of Miami College of Arts and Sciences View all articles by this author Isabella M. Olensky University of Miami College of Arts and Sciences View all articles by this author Elizabeth White University of Miami College of Arts and Sciences View all articles by this author Millie E. Rogers University of Miami College of Arts and Sciences View all articles by this author Christopher A Searcy 0009-0004-1026-4015 [email protected] University of Miami College of Arts and Sciences View all articles by this author Metrics & Citations Metrics Article Usage 476 views 389 downloads .FvxKWukQNSOunydq8rnd { width: 100px; } Citations Download citation Stephanie L. Clements, F. Michael Ackerman, Isabella M. Olensky, et al. Non-native herpetofauna continue to proliferate in the world’s most invaded herpetofauna community: Evidence against community saturation. Authorea . 17 January 2025. DOI: https://doi.org/10.22541/au.173710805.55043337/v1 If you have the appropriate software installed, you can download article citation data to the citation manager of your choice. Simply select your manager software from the list below and click Download. For more information or tips please see 'Downloading to a citation manager' in the Help menu . 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