Global economic costs of herpetofauna invasions

doi:10.21203/rs.3.rs-964112/v1

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Global economic costs of herpetofauna invasions

https://doi.org/10.21203/rs.3.rs-964112/v1

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Biological invasions by amphibian and reptile species (i.e. herpetofauna) are numerous and widespread and have caused severe impacts on ecosystems, the economy and human health. However, there remains no synthesised assessment of the economic costs of these invasions. Therefore, using the most comprehensive database on the economic costs of invasive alien species worldwide (InvaCost), we analyse the costs caused by invasive alien herpetofauna according to taxonomic, geographic, sectoral and temporal dimensions, as well as the types of these costs. The cost of invasive herpetofauna totaled 17.0 billion US$ between 1986 and 2020, divided between 6.3 billion US$ for amphibians, 10.4 billion US$ for reptiles and 334 million US$ for mixed classes. However, these costs were associated predominantly with only two species (brown tree snake Boiga irregularis and American bullfrog Lithobates catesbeianus), with 10.3 and 6.0 billion US$ in costs, respectively. Costs for the remaining 19 reported species were relatively minor (< 0.6 billion US$), and they were entirely unavailable for over 94% of known invasive herpetofauna worldwide. Also, costs were positively correlated with research effort, suggesting research biases towards well-known taxa. So far, costs have been dominated by predictions and extrapolations (79%), and thus empirical observations for impact were relatively scarce. The activity sector most affected by amphibians was authorities-stakeholders through management (> 99%), while for reptiles, impact was reported mostly through damages to mixed sectors (65%). Geographically, Oceania and Pacific Islands recorded 63% of total costs, followed by Europe (35%) and North America (2%). Cost reports have generally increased over time. A greater effort in studying the costs of invasive herpetofauna is necessary for a more complete understanding of invasion impacts of these species. We emphasise the need for greater control and prevention policies concerning the spread of current and future invasive herpetofauna.

Aquaculture and Mariculture

Biological Chemistry

Amphibia

biological invasion

economic impact

pathways

Reptilia

vertebrates

The transportation rate of alien species into new regions is unprecedented and accelerating with globalising trade and transport networks (Seebens et al., 2017). These often lead to invasions, with ecological impacts that include extinctions of native species and disruptions of ecosystem functioning (Bellard et al., 2016) through direct and indirect effects (Walsh et al., 2016; Linders et al., 2019). They also have major impacts on economic activities and human society — with high monetary costs to multiple economic sectors (Diagne et al., 2021a), disruptions of livelihoods and loss of human health (Schaffner et al., 2013, 2020) and welfare (Shackleton et al., 2019). Synthesised assessments of the impacts of major taxonomic groups are sparse (Clavero and García-Berthou 2005; Crystal-Ornelas and Lockwood 2020), inhibiting management actions, assessment of biases and the filling of knowledge gaps (Florencio et al., 2019).

In recent decades, invasive herpetofauna (i.e. amphibians and reptiles), have gained great interest on various platforms (e.g. social media) due to the numerous impacts they cause and the interest they generate from a wide range of sectors of society (Measey et al., 2019). These impacts are present in different forms, such as ecological (i.e. extirpation or reduction of native species, e.g. decline of rare species via predation), evolutionary (i.e. genetic or morphological changes, e.g. hybridisation with natives species), epidemiological (i.e. spread of disease, e.g. chytridiomycosis caused by Batrachochytrium dendrobatidis and Batrachochytrium salamandrivorans; Ossiboff et al., 2019) and socio-economic (i.e. monetary and social impacts, e.g. through agricultural productivity declines or healthcare costs; Kraus 2009a, 2015) impacts. For example, the yellow-bellied slider Trachemys scripta elegans, has degraded the functioning of ecosystems by altering abiotic characteristics, and through the displacement of native species (Ramsay et al., 2007; Lindsay et al., 2013); the cane toad Rhinella marina has led to morphological changes in the heads of native ranivorous snakes (Phillips and Shine 2006), declines in native predators due to toxicity (Shanmuganathan et al., 2010), and caused substantial management expenditures (Smart et al., 2020); the common coquí frog Eleutherodactylus coqui has led to declines in property values in areas infested due to its extremely loud mating song (Reaser et al., 2007); the brown tree snake Boiga irregularis has generated significant economic costs through island-wide power outages, resulting in a loss of US$ 4.5 million US$ per year (Fritts 2002), as well as ecological impacts through the extirpation of most endemic birds via depredation (Rodda et al., 1997).

Yet, there are now more than 1,000 known occurrences of alien reptile and amphibian species worldwide (Kraus 2011) and the number of alien herpetofauna species has continued to increase in recent decades (Kraus 2009b; Seebens et al., 2017), primarily due to the globalisation of human activity (Capinha et al., 2020) and other cultural or social factors (Reed and Kraus 2010; Wasserman et al., 2018). Thus far, the primary introduction pathways for alien herpetofauna arise from unintentional and intentional transportation by humans, for example, via stowaways in cargo shipments or the nursery trade, and via the pet trade, religious releases, aquaculture or biocontrol programmes (Kraus 2009b, 2011). Furthermore, climate change is expected to increase the invasibility of regions such as Central and South America and central Africa to alien herpetofauna (Li et al., 2016).

Knowledge gaps regarding invasive herpetofauna are pervasive due to the lack of adequate data and the presence of geographic and taxonomic biases (Bellard and Jeschke 2016). These gaps in knowledge lead to underestimations of broadscale negative impacts of these species, which creates a disincentive for invasion pathway regulatory policy and investment in control or biosecurity measures (e.g., trade of alien pets). In this context, the monetisation of the effects of invaders is a valuable option to raise societal awareness as well as motivate and justify policy action by decision-makers (Diagne et al., 2021a). The recent publication of the InvaCost database, where literature on the costs of alien species is compiled, standardised and described across a large number of indicators in a public form (Diagne et al., 2020b), has increased interest in the synthesised study of invasive alien species monetary costs (Diagne et al., 2021d; Cuthbert et al., 2021a; b; Haubrock et al., 2021a).

Here, we quantify and synthesise, for the first time, the economic costs of alien herpetofauna worldwide, based on the current state-of-the-art in the published literature. We expect our results to reveal high costs, but also patterns and possible biases that will inform the direction of future research: i) we expect a minority of species will occupy the majority of the research effort and costs (Van Wilgen et al., 2018; Cuthbert et al., 2021c; Heringer et al., 2021); ii) costs will be biased towards particular regions, as shown previously by Van Wilgen et al., (2018) and Diagne et al., 2021a); iii) most economic costs will come from damage to primary sectors such as agriculture, due to these being commonly impacted by invasions (Paini et al., 2016); and iv) temporal trends will be characterised by substantially increasing costs over time, due to the increase in invasion rates worldwide (Seebens et al., 2017) and a growing number of pathways introducing herpetofauna.

Data collection

To estimate the costs of invasive alien reptiles and amphibians worldwide, we used the most recent version (4.0; released in June 2021) of the InvaCost database (https://doi.org/10.6084/m9.figshare.12668570), per the “getInvaCostVersion” function of the invacost R package (Diagne et al., 2020b; Leroy et al., 2020; R Core Team 2021). This version of the database contains a total of 13,123 cost entries, extracted primarily from four different sources: Web of Science (https://webofknowledge.com/), Google Scholar (https://scholar.google.com/), Google search engine (https://www.google.com/) and biological invasion experts, that were contacted as part of this collection and identification process, to gather additional data. The collection and collation of data from these sources were based upon systematic search strings used to identify the costs of biological invasions (Diagne et al., 2020a). The InvaCost database contains entries reported in 15 languages in a sufficiently detailed manner to analyse the cost of invasive alien species (Angulo et al., 2021a).

Presently, InvaCost is the most complete public database on the economic costs of invasive alien species and whose methodology allows standardising costs over time and across countries. This database is continuously updated either by corrections (e.g. duplicate entries) or by adding new data. Each entry was converted into 2017 US dollars to standardise the cost values, making costs directly comparable. In addition, each entry has 65 descriptor variables, including, for example, information on species’ taxonomy, the type of ecosystem invaded and the type of economic impact caused.

Data processing

We performed a series of steps before proceeding with the analysis of the data (Figure 1). First, we refined the database by eliminating data with no information on starting and ending years (i.e. incomplete temporal scale; see later) and no monetised cost in the column “Cost_estimate_per_year_2017_USD_exchange_rate”. Next, we filtered the database to include only the “Amphibia” and “Reptilia” classes (column: “Class”, n = 354). Entries belonging to the class “Amphibia/Reptilia” and “Reptilia/Amphibia” were included in a new “Class” group (“Diverse”, n = 6). On the other hand, an entry recorded as “Amphibia/Arthropoda/Mammalia/Diverse” was excluded from the study due to the impossibility of separating the unique costs of amphibians. The total costs of the alien species were calculated with the function “summarizeCosts” of the invacost package (Leroy et al., 2020).

This final dataset contained 353 entries which were all homogenised and converted on an annual basis, whereby the total cost for each estimate and each species was divided by the duration time of each cost entry. This process of ‘expanding’ the dataset was performed by the function “expandYearlyCosts'' of the invacost R package (Leroy et al., 2020; R Core Team 2021). This function expands the estimated costs over the years they occurred or were estimated to have occurred. This means that one cost entry that reports a total cost of 100 million over 10 years, for example, will be transformed to ten cost entries with an annual cost of 10 million dollars each. The duration of the study was determined from the columns “Probable_starting_year_adjusted” for the start of the impact and “Probable_ending_year_adjusted” for the end of the impact. All analyses were carried out on this expanded database to reflect the duration of the costs according to each entry. Our final, expanded dataset had 710 entries from 1986 to 2020, thus covering 35 years.

Cost descriptors

The cumulative costs of invasive alien species were aggregated according to the following descriptor variables:

Method reliability: Parameter used to indicate the degree of reliability of the estimation of costs. A cost is classified as High reliability if it is reported from official reports or peer-reviewed articles, and if not, whether the estimation method is reproducible; if these criteria are not met, the estimate will be categorised as being of Low reliability.

Implementation: Indicates whether the cost estimate has been produced directly in the study area (Observed), or whether it is instead a predicted cost derived from the distribution area of alien species elsewhere, i.e., a cost not incurred but expected (Potential).

Species: This descriptor shows the cost totals per species. In addition, entries that contained more than one species that could not be distinguished (column: “Species”) were added to the “Species” category “Diverse” (n = 44).

Impacted sector: market and/or activity sector affected by the costs of invasive herpetofauna, (e.g. Agriculture, Health, Public and social welfare). All entries with more than one category were pooled into “Diverse”.

Type of costs: Type of cost generated by the invasion: Damage (e.g. yield loss, health injury), Management (e.g. control, research) and Mixed (indistinguishable damage and management components).

Geographic region: Refers to the continental region where the cost of alien species invasion was recorded. We also made several changes to the descriptor variable “Geographic_region” to homogenise the sub-group. These changes were the joining of “Pacific Islands” and “Pacific Islands/Oceania” into “Oceania and Pacific Islands”, and “Central America” and “South America” into “Central/South America” and “Oceania/South America'' into ''Diverse”.

We also examined the proportion of invasive alien species for which costs were reported using a database reporting alien herpetofauna worldwide, separately by order and family (Capinha et al., 2017). To our knowledge, this is the most complete database published on invasive herpetofauna records, whereby each entry represents the occurrence of an alien established species in a country, federal state or biogeographically-separate island or archipelago.

Furthermore, we examined whether species with costs were more studied than those without costs, and correlated the magnitude of costs with the numbers of studies. A search in the Web of Science (Appendix A). We used an unpaired two-sample Wilcoxon test with an alpha level of 0.05 to compare study numbers between species with (i.e., in InvaCost) and without (i.e., not in Invacost) costs. We then assessed whether the magnitude of total and observed costs of invasive alien species in InvaCost was correlated to their respective numbers of associated papers through linear regressions. We assessed the normality of residuals using a Shapiro-Wilks test and performed a logarithmic transformation [log10(x + 1)], as costs did not comply with this assumption to normalise the variables.

To assess the trend of alien reptilian and amphibian species costs, we considered 5-year intervals ranging from 1986 to 2020 due to data availability. We used the “summarizeCosts” function of the invacost package (Leroy et al., 2020) to estimate the total costs of alien species per year and the mean for each five-year period, considering amphibians and reptiles separately.

Economic costs by order and family

The total reported costs of the invasive herpetofauna exceeded 16.98 billion US$ between 1986 to 2020. We found that out of 280 invasive herpetofauna species recorded in Capinha et al. (2017), 263 species did not have any economic impacts recorded in the InvaCost database (93.92 %, Table S1). For amphibians, we found that 75 of 81 species did not record a cost (92.59 %), while 188 of 199 reptiles did not record a cost (94.47 %). In particular, no cost was recorded for Caudata (Amphibia) and Crocodilia (Reptilia) entirely (Table S1, Appendix B). This lack of studies is magnified at the family level, where we only found costs for 5 out of 20 (25 %) families of amphibians and 7 out of 37 (18.91 %) families of reptiles (Table S1). In addition, in Hylidae and Salamandridae, even though the most biodiverse families in terms of alien amphibian species from Capinha et al. (2017), no economic cost was reported. Similarly, for the richest families of reptiles, Gekkonidae or Scincidae, no costs were reported.

Cost among classes, reliability and implementation type

While amphibians contributed a total of 6.27 billion US$, nearly all costs came from high reliability entries, amounting to more than 6.24 billion US$ (99.58%), compared to 26.08 million US$ (0.42%) in low reliability entries. In terms of implementation, only 70.56 million US$ came from observed costs (1.28% of costs), and 6.19 billion US$ came from potential costs (98.72% of costs). For reptiles, 8.12 billion US$ (78.10%) of reptile cost entries came from high reliability sources and 2.25 billion US$ (21.89%) came from low reliability sources (Figure 2). However, a larger share of 3.49 billion US$ (33.65%) of the total 10.37 billion US$ came from observed costs, while 6.88 billion US$ (66.35%) came from potential costs. For the diverse (i.e., mixed) class, all recorded costs (334.14 million US$) from diverse classes came from high reliability sources (Figure 2). In addition, we found a total cost of 334.14 million US$ (1.96% of total herps costs), whereby 327.01 million US$ (97.90%) came from potential costs and 7.13 million US$ came from observed costs (2.10%)

Cost among species and taxonomic group

The overall cost of invasive herpetofauna was due to a handful of species, which contributed the bulk of total costs (Figure 3). For amphibians, the total cost of 6.27 billion US$ was divided among only six species, but a very large part of these costs came from a single species (American bullfrog, Lithobates catesbeianus) with a total cost of 6.04 billion US$ (96.33 %, n = 63). The rest of the amphibian species accumulated less than 0.23 billion US$ (3.66 %. n = 293).

For the reptiles, the total cost of 10.37 billion US$ was divided among 15 species, with the B. irregularis registering a total of 10.34 billion US$ (99.30 %, n = 78), while the rest of reptilian species accumulated less than 0.03 billion US$ (0.70 %). The species with the highest number of entries were T. scripta (n = 109), B. irregularis (n = 78) and the Green anole Anolis carolinensis (n = 60); the remaining species accounted for 91 entries in total.

For the amphibian, most of the observed costs belonged to R. marina (43.5 million US$; 61.42 %; n = 155), followed by E. coqui with 18.30 million US$ (26.14 %; n = 6). The remaining amphibian species (n = 5) recorded less than 8.3 million US$ (< 12.50 %; n = 93).

For the reptiles, observed species-level costs totalled 3.49 billion US$, and B. irregularis dominated with 3.46 billion US$ (99.14 %; n = 38), while the remaining species (n = 14) cumulated less than 28 million US$ (< 0.90 %; n = 254).

We identified significantly greater (W = 4051.5, p-value = < 0.001) numbers of publications for the species with recorded costs in the InvaCost database compared to those that did not have any record of economic costs. Further, for those species with costs, we found a significant positive correlation between the total cost of the recorded species and the number of respective publications (Adjusted R² = 0.26, p-value = 0.01) (Figure S1, Appendix B). This relationship was also significantly positive in the case of observed costs only (Adjusted R² = 0.16, p-value = 0.04) (Figure S2, Appendix B).

Costs among activity sectors and types

For the amphibian, we found that the Authorities-stakeholders activity sector reported the highest economic costs (6.25 billion US$, 99.6 %, n = 335) followed by Public and social welfare (0.02 billion US$, 0.24 %, n = 2) and Agriculture (8.98 million US$, 0.14 %, n = 23), while the remaining impacted sectors recorded less than 728.88 thousand US$ together (< 0.01 %; n = 6) (Figure 4). Focusing on observed costs only, we found that the Authorities-stakeholders sector was still the most affected (48.04 million US$, 68.09 %) by invasive amphibians, followed by Public and social welfare (13.27 million US$, 18.81 %) and Agriculture (8.84 million US$, 12.52 %). The remaining impacted sectors reported less than 397.12 thousand US$, <0.56%).

For the reptiles, we found that the most impacted sector was Diverse (i.e. mixed; 6.75 billion US$, 65.1 %, n = 37) followed by Public and social welfare (3.39 billion US$, 32.7 %, n = 8) and Authorities-stakeholders (0.22 billion US$, 2.17 %, n = 278). The remaining impacted sectors recorded less than 415.38 thousand US$ together (< 0.003 %, n = 15) (Figure 4). This pattern was not repeated for observed costs considering invasive reptiles, with Public and social welfare (3.39 billion US$, 97.05 %), followed by Authorities-stakeholders (99.44 million US$, 2.84 %) and then the Diverse (3.10 million US$, 0.08 %) sectors. The remaining impacted sectors reported less than 412.20 thousand US$ (<0.01 %).

With respect to type of cost for amphibians, we found that the highest costs were for Management (6.25 billion US$, 99.7 %, n = 385), followed by Damage (15.60 million US$, 0.25 %, n = 3) and Mixed (441.91 thousand US$, 0.007 %, n = 5) types (Figure 4). Focusing on Observed costs only, we found again that the highest costs were for Management (56.84 million US$, 80.55 %) followed by Damage (13.27 million US$, 18.81 %) and Mixed (441.98 thousand US$, 0.62 %) types.

For reptiles, we conversely found that the highest cost came from Damage (10.02 billion US$, 96.6%, n = 251), followed by Management (0.22 billion US$, 2.16%, n = 35) and Mixed (0.12 billion US$, 1.22 %, n = 25) types (Figure 4). This pattern was repeated for Observed costs, with first Damage (3.39 billion US$, 97.16 %), followed by Management (99.03 million US$, 2.83 %) and Mixed (43.96 thousand US$, 0.001%) types.

Economic cost across geographic regions

Reported costs were strongly geographically clustered among classes. For the amphibians, we found that Europe incurred the highest economic costs (6.04 billion US$, 96.26 %, n = 68) followed by Oceania and Pacific Islands (0.23 billion US$, 3.67 %, n = 240) and diverse regions (3.25 million US$, 0.05 %. n = 8). Of these costs, the majority was Potential (6.19 billion US$, 98.92 %), being mostly in Europe (6.04 billion US$) and little for Oceania and Pacific Islands (0.17 billion US$), while Observed costs were less than 0.08 billion US$. The regions with the highest Observed costs were Oceania and Pacific Islands (62.14 million US$, 88.07 %), followed by Europe (3.94 million US$, 5.59 %) and Diverse (3.25 million US$, 4.61 %) (Figure 5).

For the reptiles, we found that Oceania and Pacific Islands reported the highest costs (10.5 billion US$, 99.61 %, n = 77), followed by Asia (25.00 million US$, 0.24 %, n = 112) and North America (12.97 million US$, 0.12 %, n = 18). Here, 66.24 % came from Potential costs, with 6.87 billion US$ of the total, and 3.49 billion US$ (33.76 %) came from Observed costs. The regions with the most costs observed were, again, Oceania and Pacific Islands (3.46 billion US$, 98.85 %) followed by Asia (25.00 million US$, 0.71 %) and North America (12.96 million US$, 0.37 %) (Figure 5).

Temporal trends

The total cost of invasive alien herpetofauna exceeded 16.98 billion US$ globally between 1986 and 2020, averaging US$ 485.21 million per year. The average annual cost of invasive amphibians totalled 179.29 million US$, with 1.21 billion US$ being the highest average annual cost in the period from 2011 to 2015 (Figure 6a). The average annual cost of invasive reptiles was 296.37 million US$, with 1.43 billion US$, the highest average cost in the period from 2006 to 2010 (Figure 6b).

Concerning only Observed costs for the amphibian class, we found 70.56 million US$ in total for the entire period, with the average annual cost being 2.01 million US$; the period in which the total average annual cost was the highest corresponds to 2006 to 2010 with 5.87 million US$. On the other hand, for the reptiles, an Observed cost of 3.49 billion US$ was found, corresponding to an average annual cost over the entire period of 99.88 million US$, with costs peaking between 2001 to 2005 with an average annual cost of 0.45 billion US$ (Figure 6).

The present study found a total reported cost of global herpetofauna invasions of 16.98 billion US$ since 1986. However, reported costs were principally due to just a few species, namely B. irregularis, L. catesbeianus and R. marina (~96 %). Given the notoriety of these taxa in causing impacts (e.g. all three appear on the IUCN “100 of the world’s worst invasive species” list; Lowe et al., 2000; Cuthbert et al., 2021a), there could be a potential positive feedback loop between research effort and cost reporting. Indeed, we found significantly greater numbers of papers concerning species with reported costs than those without reported costs. This corroborates earlier identifications of bias in studies of invasive alien herpetofauna (Van Wilgen et al., 2018). While we acknowledge that not all invasive alien herpetofauna will have tangible monetary impacts, greater research effort is required to distinguish the true absence of economic cost from gaps in cost detection.

The cost of invasive herpetofauna exceeds costs reported from other invasive taxa using the InvaCost database, such as rodents (3.28 billion US$, Diagne et al., 2021c), freshwater crayfish (120.50 million US$, Kouba et al., 2021) and the cost of all marine species (3.6 billion US$, Cuthbert et al., 2021b). Most of the reported costs and entries were obtained from peer-reviewed literature or were deemed of high reliability. We note, however, that a binary classification of method reliability does not account for the variation of methods employed for varying taxa and regions in individual studies; they nevertheless provided a practical way of objectively assigning cost reliability. Further, these costs predominantly came from estimates and extrapolations, being evaluated as potential costs rather than empirically observed. Additionally, whereas the potential costs were higher, the number of entries was lower than the observed costs, indicating that potential cost entries tend to report much greater costs, which is to be expected since they often extrapolate costs from smaller to larger scales (e.g., from sub-national to national scales or for longer periods; Diagne et al., 2020a).

We found that most of the reported costs for amphibians are from management, while most other studies consistently show damage as the main source of costs, with values of 10 to 100 times higher than management (Diagne et al., 2021e; Bradshaw et al., 2021). This could indicate a massive underestimation of this cost type, although this pattern is not found in reptiles. Especially surprising is the low reported cost from Burmese python Python bivittatus considering the large number of impacts generated in the USA by the high predation capacity of this species, which has caused the decline of native species (Dorcas et al., 2012), as well as great effort in the country to control its populations (Mazzotti et al., 2016). Yet, many of these ecological impacts are not monetised or available as they are not in the InvaCost database, and all recorded economic costs generated by P. bivittatus thus far are from a single eradication program targeting this species in Florida (Smith et al., 2016). In addition, there is a possible cost bias as our study range is limited to 1986 - 2020, while some of the most notorious species were introduced long before that [e.g. R. marina were introduced in 1936 in Australia (Eastel 1981)].

The lack of reported costs found for the vast majority of reported alien species corroborates other studies at various spatial and taxonomic scales, including for the United Kingdom (Cuthbert et al., 2021c), Singapore and Southeast Asia (Haubrock et al., 2021c), fishes (Haubrook et al., 2021b), and ants (Angulo et al., 2021b), among others. So, some families or species concentrate most of the economic costs, and may be linked to their unique characteristics [e.g. toxicity (R. marina) or popular exotic pet trade (Colubridae)].

Concerning sectors of economic activity for amphibians, most costs impacted authorities-stakeholders. This means that most of these costs come from management programmes, such as for control and eradication of alien species, which is motivated by increasingly stringent laws [e.g., Regulation (EU) No 1143/2014 of the European Parliament and of the Council, as is the case for L. catesbeianus included in this list] and it suggests that most impacts are on non-monetizable assets, such as biodiversity. Conversely, a more diverse pattern was found for reptiles, emphasising a large amount of damage from these species across various sectors.

While taxonomic gaps were pervasive across socio-economic sectors, these were coupled with a similar lack of data for geographic regions (although there may also be bias in geographical economic costs due to factors such as language, but see Angulo et al., 2021a). The regions most studied were Oceania and Pacific Islands, Europe and Asia, while the least studied were Africa, Central America and South America. However, this difference may also arise from the number of alien species established in each region and research capacity in general. Most of the studies in Oceania and Pacific Islands corresponded to the effects of R. marina in Australia (Taylor and Edwards 2005) and B. irregularis in Hawaii and Guam (Burnett et al., 2012). Surprisingly, and in contrast to other studies (Cuthbert et al., 2021b; Haubrok et al., 2021a), a small number of studies from North America were observed. North American (especially in the USA) studies pioneered early works in economic costs estimations for invasive alien species (Pimentel et al., 2005), promoting further research in the last few decades that resulted in burgeoning reported economic impacts from invasions there (Fantle-Lepczyk, et al., 2021; Crystal-Ornelas et al., 2021).

Between the two classes, we found that the bulk of the costs were recorded from reptiles, especially in Oceania and Pacific Islands, except for Europe where amphibian costs were dominant for both total and observed costs. In particular, a large part of these European costs stems from the extrapolation of the economic impact of L. catesbeianus in Germany (European Environment Agency 2013). For amphibians, no costs were recorded in Africa, although there are several species of invasive alien amphibians in this region that cause problems, such as guttural toad Sclerophrys gutturalis, generating direct impact on the local economy by the displacement of native species and species eradication programmes (Measey et al., 2017). For reptiles, contrary to what has been found with amphibians, Europe was one of the regions with the lowest recorded costs, alongside Central and South America. However, this could change in future years since, as demonstrated by Li et al. (2016), the Central and South American region is expected to be most affected by herpetofauna invasions.

In addition, costs from B. irregularis here were centred on studies from the Pacific islands (Hawaii, Guam and northern Mariana Islands). Island ecosystems are regarded as more vulnerable to invasion by alien species and their impacts (Anton et al., 2020). They can incur economic costs through a different suite of sectors and species than mainland areas (Rico-Sánchez et al., 2021). Islands are substantially more biodiverse than mainlands (McNeely 2001). While ecological impacts can be challenging to monetise, alien species there generate high rates of extinction of native species in groups such as vertebrates and plants (Sax and Gaines 2008), which can have high intrinsic or cultural value (Christie et al., 2012). Nonetheless, our study may help develop more effective policies to reduce these future economic costs.

Analysis of the cost trend between 1986 to 2020 revealed similar patterns for both classes. For amphibians, the cost increased until 2006, when it declined steeply. For the reptile, the cost increased until 2006-2010, when it declined steeply until the final years of this study. Primarily, these trends likely relate to the delay in reporting impacts (Diagne et al., 2021a), as well as the possibility that there is a lag phase in the impacts generated by these species, which can take decades (Essl et al., 2011). Indeed, the range in which the greatest costs of amphibians was detected coincides with the years with the highest number of entries, with a total of 105 entries between 2006 to 2010, while recent years' entries record a much lower amount, with only 46 entries between 2016 to 2020. This pattern was not seen in reptiles, where in the period with the highest number of costs in 2011 to 2015, only 40 entries were reported. On the contrary, 149 entries were reported in 2016-2020. We therefore suggest that the presented costs over time for herpetofauna invasions have been heavily influenced by the evolution of the costs of the three costliest species (B. irregularis, L. catesbeianus and R. marina) and the timing of underlying studies.

This study presents the first global synthesis of the costs of invasive herpetofauna based on the current state-of-the-art in the InvaCost database. While costs are high, they are likely severely underestimated due to knowledge gaps, with costs focused on only a few taxa, regions and limited to recent decades (see also Diagne et al. 2021d for methodological and philosophical underlying reasons). Furthermore, because invasion rates are expected to increase in the future (Seebens et al., 2021), we expect this increase to be followed by a rise in economic costs. Yet, it could be reduced by investing in management, especially in the prevention of the arrival of alien species and early detection measures (Ahmed et al., 2021; Leung et al., 2002).

Further, the main pathway of invasion of alien herpetofauna is the pet trade (Kopecký et al., 2013). For amphibians, it is expected that future trade will not differ from the current patterning, where some of the species that are expected to still be traded widely are the yellow cururu toad Rhinella icterica, Müller's platanna Xenopus muelleri or Woodhouse's toad Anaxyrus woodhousii (Mohanty and Measey 2019). In contrast, the reptile pet trade is expected to change its most traded species in the future, expectedly resulting in new economic impacts to potentially different sectors. While I. iguana, P. bivittatus and T. scripta have declined the most in popularity, other species such as blue-tongued lizards Tiliqua spp., crested geckos Correlophus ciliatus, and ball pythons Python regius are expected to increase (Valdez 2021).

Economic impacts of invasive herpetofauna are a noteworthy and global problem, yet we found information to be spatially fragmented. Furthermore, the majority of the recorded costs originated from few species, a better understanding of the costs of invasive amphibians and reptiles can contribute to a higher awareness of the risk of introductions, leading to more effective regulatory instruments. Thus, we highlight the urgent need to further quantify costs incurred by invasive herpetofauna given the high context-dependency in invasion science and recommend conducting assessments at more granular spatial and taxonomic scales (Ricciardi 2013). Ultimately, we emphasize the importance of the development of policies by authorities and governments to prevent the arrival of alien species, since effective prevention will decrease future expenses (Ahmed et al., 2021).

Author contributions

I.S. designed of the work, data elaboration and interpretation, R.C. design of the work, interpretation of data and have drafted the work and substantively revised it, A.K. interpretation of data and substantively revised it, C.C. interpretation of data and substantively revised it, A.T. substantively revised it, E.H. interpretation of data and substantively revised it, C.D. interpretation of data and substantively revised it, F.C. acquisition of the database, interpretation of data and substantively revised it, P.H. conception, designed of the work, interpretation of the data and have drafted the work, and substantively revised it. All authors reviewed the manuscript.

Acknowledgements

The authors acknowledge the French National Research Agency (ANR-14-CE02-0021) and the BNP-Paribas Foundation Climate Initiative for funding the InvaCost project that allowed the construction of the InvaCost database. The present work was conducted following a workshop funded by the AXA Research Fund Chair of Invasion Biology and is part of the AlienScenarios project funded by BiodivERsA and Belmont-Forum call 2018 on biodiversity scenarios. C.C. was supported by Portuguese National Funds through FCT, I.P., under the programme of “Stimulus of Scientific Employment—Individual Support” within the contracts “CEECIND/02037/2017”. RNC acknowledges funding from the Alexander von Humboldt Foundation.

Additional information

The authors declare no competing interests.

Correspondence and requests for materials should be addressed to P.H.

Ahmed, D. A., Hudgins, E. J., Cuthbert, R. N., Kourantidou, M., Diagne, C., Haubrock, P. J., Leung, B., Liu, C., Leroy, B., & Courchamp, F. (2021). Managing biological invasions: the cost of inaction.

Angulo, E., Diagne, C., Ballesteros-Mejia, L., Adamjy, T., Ahmed, D. A., Akulov, E., Banerjee, A., Capinha, C., Dia, C., Dobigny, G., Duboscq-Carra, V., Golivets, M., Haubrok, P.J., Heringer, G., Kirichencko, N., Kourantidou, M., Liu, C., Nuñez, M., Renault, D., & Courchamp, F. (2021a). Non-English languages enrich scientific knowledge: The example of economic costs of biological invasions. Science of the Total Environment, 775, 144441.

Angulo, E., Hoffmann, B., Ballesteros-Mejia, L., Taheri, A., Balzani, P., Renault, D., Cordonnier, M., Bellard, C., Diagne, C., Ahmed, D. A., Watari, Y., & Courchamp, F. (2021b). Economic costs of invasive alien ants worldwide.

Anton, A., Geraldi, N. R., Ricciardi, A., & Dick, J. T. (2020). Global determinants of prey naiveté to exotic predators. Proceedings of the Royal Society B, 287(1928), 20192978.

Bellard, C., Cassey, P., & Blackburn, T. M. (2016). Alien species as a driver of recent extinctions. Biology letters, 12(2), 20150623.

Bellard, C., & Jeschke, J. M. (2016). A spatial mismatch between invader impacts and research publications. Conservation Biology, 30(1), 230-232.

Bradshaw, C. J., Hoskins, A. J., Haubrock, P. J., Cuthbert, R. N., Diagne, C., Leroy, B., Andrews, L., Page, B., Cassey, P., Sheppard, A. W., & Courchamp, F. (2021). Detailed assessment of the reported economic costs of invasive species in Australia. NeoBiota, 67, 511-550.

Burnett, K., Pongkijvorasin, S., & Roumasset, J. (2012). Species invasion as catastrophe: the case of the brown tree snake. Environmental and Resource Economics, 51(2), 241-254.

Capinha, C., Marcolin, F., & Reino, L. (2020). Human‐induced globalization of insular herpetofaunas. Global Ecology and Biogeography, 29(8), 1328-1349.

Capinha, C., Seebens, H., Cassey, P., García‐Díaz, P., Lenzner, B., Mang, T., Moser, D., Pyšek, P., Rödder, D., Scalera, R., Winter, M., Dullinger, S., & Essl, F. (2017). Diversity, biogeography and the global flows of alien amphibians and reptiles. Diversity and Distributions, 23(11), 1313-1322.

Christie, M., Fazey, I., Cooper, R., Hyde, T., & Kenter, J. O. (2012). An evaluation of monetary and non-monetary techniques for assessing the importance of biodiversity and ecosystem services to people in countries with developing economies. Ecological economics, 83, 67-78.

Clavero, M., & García-Berthou, E. (2005). Invasive species are a leading cause of animal extinctions. Trends in ecology & evolution, 20(3), 110.

Crystal-Ornelas, R., Hudgins, E. J., Cuthbert, R. N., Haubrock, P. J., Fantle-Lepczyk, J., Angulo, E., Kramer, A. M., Ballesteros-Mejia, L., Leroy, B., Leung, B., Lopez-Lopez, E., Diagne, C., & Courchamp, F. (2021). Economic costs of biological invasions within North America. NeoBiota, 67, 485.

Crystal-Ornelas, R., & Lockwood, J. L. (2020). The ‘known unknowns’ of invasive species impact measurement. Biological Invasions, 22(4), 1513-1525.

Cuthbert, R. N., Diagne, C., Haubrock, P. J., Turbelin, A. J., & Courchamp, F. (2021a). Are the “100 of the world’s worst” invasive species also the costliest?. Biological Invasions, 1-10.

Cuthbert, R. N., Pattison, Z., Taylor, N. G., Verbrugge, L., Diagne, C., Ahmed, D. A., Leroy, B., Angulo, E., Briski, E., Capinha, C., Catford, J., Dalu, T., Essl, F., Gozlan, R., Haubrock, P. J., Kourantidou, M., Kramer, A. M., Renault, D., Wasserman, R., & Courchamp, F. (2021b). Global economic costs of aquatic invasive alien species. Science of the Total Environment, 775, 145238.

Cuthbert, R., Bartlett, A., Turbelin, A., Haubrock, P., Diagne, C., Pattison, Z., Courchamp, F., & Catford, J. (2021c). Economic costs of biological invasions in the United Kingdom. Neobiota.

Diagne, C., Ballesteros-Mejia, L., Bodey, T., Cuthbert, R., Fantle-Lepczyk, J., Angulo, E., Dobigny, G., & Courchamp, F. (2021c). Economic costs of invasive rodents worldwide: the tip of the iceberg.

Diagne, C., Catford, J. A., Essl, F., Nuñez, M. A., & Courchamp, F. (2020a). What are the economic costs of biological invasions? A complex topic requiring international and interdisciplinary expertise. NeoBiota, 63, 25.

Diagne, C., Leroy, B., Gozlan, R. E., Vaissière, A. C., Assailly, C., Nuninger, L., Roiz, D., Jourdain, F., Jarić, I., & Courchamp, F. (2020b). InvaCost, a public database of the economic costs of biological invasions worldwide. Scientific data, 7(1), 1-12.

Diagne, C., Leroy, B., Vaissière, A. C., Gozlan, R. E., Roiz, D., Jarić, I., Salles, J. M., Corey, J., Bradshaw, C., & Courchamp, F. (2021e). High and rising economic costs of biological invasions worldwide. Nature, 592(7855), 571-576.

Diagne, C., Turbelin, A., Moodley, D., Novoa, A., Leroy, B., Angulo, E., Adamjy, T., Dia, C., Ahmed, T., Tambo, J., Dobigny, G., & Courchamp, F. (2021d). The economic costs of biological invasions in Africa: a growing but neglected threat?. NeoBiota, 67, 11-51.

Dorcas, M. E., Willson, J. D., Reed, R. N., Snow, R. W., Rochford, M. R., Miller, M. A., Meshaka, W., Andreadis, P., Mazzotti, F., Romagosa, C., & Hart, K. M. (2012). Severe mammal declines coincide with proliferation of invasive Burmese pythons in Everglades National Park. Proceedings of the National Academy of Sciences, 109(7), 2418-2422.

Easteal, S. (1981). The history of introductions of Bufo marinus (Amphibia: Anura); a natural experiment in evolution. Biological Journal of the Linnean Society, 16(2), 93-113.

Essl, F., Dullinger, S., Rabitsch, W., Hulme, P. E., Hülber, K., Jarošík, V., Kleinbauer, I., Krausmann, F., Kühn, I., Nentwig, W.,, Vilá, M., Genovesi, P., Gherardi, F., Desprez-Loustau, M. L., Roques, A., & Pyšek, P. (2011). Socioeconomic legacy yields an invasion debt. Proceedings of the National Academy of Sciences, 108(1), 203-207.

European Environment Agency. (2013). The impacts of invasive alien species in Europe. Publications Office of the European Union.

Fantle-Lepczyk, J. E., Haubrock, P. J., Kramer, A. M., Cuthbert, R. N., Turbelin, A. J., Crystal-Ornelas, R., Diagne, C., & Courchamp, F. (2021). Economic costs of biological invasions in the United States. bioRxiv.

Florencio, M., Lobo, J. M., & Bini, L. M. (2019). Biases in global effects of exotic species on local invertebrates: a systematic review. Biological Invasions, 21(10), 3043-3061.

Fritts, T. H. (2002). Economic costs of electrical system instability and power outages caused by snakes on the island of Guam. International biodeterioration & biodegradation, 49(2-3), 93-100.

Haubrock, P. J., Bernery, C., Cuthbert, R. N., Liu, C., Kourantidou, M., Leroy, B., Turbelin, A., Kramer, A. M., Verbrugge, L., Diagne, C., Courchamp, F., & Gozlan, R. E. (2021b). What is the recorded economic cost of alien invasive fishes worldwide?.

Haubrock, P. J., Cuthbert, R. N., Ricciardi, A., Diagne, C., & Courchamp, F. (2021a). Massive economic costs of invasive bivalves in freshwater ecosystems.

Haubrock, P. J., Cuthbert, R. N., Yeo, D. C., Banerjee, A. K., Liu, C., Diagne, C., & Courchamp, F. (2021c). Biological invasions in Singapore and Southeast Asia: data gaps fail to mask potentially massive economic costs. NeoBiota, 67, 131-152.

Heringer, G., Angulo, E., Ballesteros-Mejia, L., Capinha, C., Courchamp, F., Diagne, C., Duboscq-Carra, V. G., Nuñez, M. A., & Zenni, R. D. (2021). The economic costs of biological invasions in Central and South America: a first regional assessment. NeoBiota, 67, 401.

Kopecký, O., Kalous, L., & Patoka, J. (2013). Establishment risk from pet-trade freshwater turtles in the European Union. Knowledge and Management of Aquatic Ecosystems, (410), 02.

Kouba, A ,. Oficialdegui, F ,. Cuthbert, R., Kourantidou, M., South., Tricarico, E ., Leroy, B., Gozlan, R., Courchamp, F., & Haubrock, P. (2021). Feeling the pinch: global economic costs of crayfish invasions and comparison with other aquatic crustaceans.

Kraus, F. (2009a). Alien reptiles and amphibians: a scientific compendium and analysis (Vol. 4). Springer Science & Business Media.

Kraus, F. (2009b). Global trends in alien reptiles and amphibians. Aliens: The Invasive Species Bulletin, 28, 13-18.

Kraus, F. (2011). Reptiles and amphibians. In Encyclopedia of biological invasions (pp. 590-594). University of California Press.

Kraus, F. (2015). Impacts from invasive reptiles and amphibians. Annual Review of Ecology, Evolution, and Systematics, 46, 75-97.

Leroy, B., Kramer, A. M., Vaissière, A. C., Courchamp, F., & Diagne, C. (2020). Analysing global economic costs of invasive alien species with the invacost R package. bioRxiv.

Leung, B., Lodge, D. M., Finnoff, D., Shogren, J. F., Lewis, M. A., & Lamberti, G. (2002). An ounce of prevention or a pound of cure: bioeconomic risk analysis of invasive species. Proceedings of the Royal Society of London. Series B: Biological Sciences, 269(1508), 2407-2413.

Li, X., Liu, X., Kraus, F., Tingley, R., & Li, Y. (2016). Risk of biological invasions is concentrated in biodiversity hotspots. Frontiers in Ecology and the Environment, 14(8), 411-417.

Linders, T. E. W., Schaffner, U., Eschen, R., Abebe, A., Choge, S. K., Nigatu, L., Mbaabu, P. R., Shiferaw, H., & Allan, E. (2019). Direct and indirect effects of invasive species: Biodiversity loss is a major mechanism by which an invasive tree affects ecosystem functioning. Journal of Ecology, 107(6), 2660-2672.

Lindsay, M. K., Zhang, Y., Forstner, M. R., & Hahn, D. (2013). Effects of the freshwater turtle Trachemys scripta elegans on ecosystem functioning: an approach in experimental ponds. Amphibia-Reptilia, 34(1), 75-84.

Lowe, S., Browne, M., Boudjelas, S., & De Poorter, M. (2000). 100 of the world's worst invasive alien species: a selection from the global invasive species database (Vol. 12). Auckland: Invasive Species Specialist Group.

Mazzotti, F. J., Rochford, M., Vinci, J., Jeffery, B. M., Eckles, J. K., Dove, C., & Sommers, K. P. (2016). Implications of the 2013 Python Challenge® for ecology and management of Python molorus bivittatus (Burmese Python) in Florida. Southeastern Naturalist, 15(sp8), 63-74.

McNeely, J. (2001). Invasive species: a costly catastrophe for native biodiversity. Land Use and Water Resources Research, 1(1732-2016-140260).

Measey, J., Davies, S. J., Vimercati, G., Rebelo, A., Schmidt, W., & Turner, A. (2017). Invasive amphibians in southern Africa: a review of invasion pathways. Bothalia-African Biodiversity & Conservation, 47(2), 1-12.

Measey, J., Basson, A., Rebelo, A. D., Nunes, A. L., Vimercati, G., Louw, M., & Mohanty, N. P. (2019). Why have a pet amphibian? Insights from YouTube. Frontiers in Ecology and Evolution, 7, 52.

Mohanty, N. P., & Measey, J. (2019). The global pet trade in amphibians: species traits, taxonomic bias, and future directions. Biodiversity and Conservation, 28(14), 3915-3923.

Ossiboff, R. J., Towe, A. E., Brown, M. A., Longo, A. V., Lips, K. R., Miller, D. L., Carter, E., Gray, M., & Frasca Jr, S. (2019). Differentiating Batrachochytrium dendrobatidis and B. salamandrivorans in amphibian chytridiomycosis using RNAScope® in situ hybridization. Frontiers in Veterinary Science, 6, 304.

Paini, D. R., Sheppard, A. W., Cook, D. C., De Barro, P. J., Worner, S. P., & Thomas, M. B. (2016). Global threat to agriculture from invasive species. Proceedings of the National Academy of Sciences, 113(27), 7575-7579.

Phillips, B. L., & Shine, R. (2006). An invasive species induces rapid adaptive change in a native predator: cane toads and black snakes in Australia. Proceedings of the Royal Society B: Biological Sciences, 273(1593), 1545-1550.

Pimentel, D., Zuniga, R., & Morrison, D. (2005). Update on the environmental and economic costs associated with alien-invasive species in the United States. Ecological economics, 52(3), 273-288.

Ramsay, N. F., Ng, P. K. A., O'Riordan, R. M., & Chou, L. M. (2007). The red-eared slider (Trachemys scripta elegans) in Asia: a review. Biological invaders in inland waters: profiles, distribution, and threats, 161-174.

R Core Team (2021). R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria.

Reaser, J. K., Meyerson, L. A., Cronk, Q., De Poorter, M. A. J., Eldrege, L. G., Green, E., Kairo, M. T. K., Latasi, P., Mack, R., Mauremootoo, J., O'Dowd, D., Orapa, W., Sastroutomo, S., Saunders, A. S., Thrainsson, S., & Vaiutu, L. (2007). Ecological and socioeconomic impacts of invasive alien species in island ecosystems. Environmental Conservation, 98-111.

Reed, R. N., & Kraus, F. (2010). Invasive reptiles and amphibians: global perspectives and local solutions. Animal Conservation, 13(Supplement 1), 3-4.

Ricciardi, A. (2013). Invasive species. In Ecological systems (pp. 161-178). Springer, New York, NY

Rico-Sánchez, A. E., Haubrock, P. J., Cuthbert, R. N., Angulo, E., Ballesteros-Mejia, L., López-López, E., Duboscq-Carra, V. G., Nuñez, M. A., Diagne, C., & Courchamp, F. (2021). Economic costs of invasive alien species in Mexico. NeoBiota, 67, 459-483.

Rodda, G. H., Fritts, T. H., & Chiszar, D. (1997). The disappearance of Guam's wildlife. BioScience, 47(9), 565-574.

Sax, D. F., & Gaines, S. D. (2008). Species invasions and extinction: the future of native biodiversity on islands. Proceedings of the National Academy of Sciences, 105(Supplement 1), 11490-11497.

Schaffner, F., Medlock, J. M., & Van Bortel, A. W. (2013). Public health significance of invasive mosquitoes in Europe. Clinical microbiology and infection, 19(8), 685-692.

Schaffner, U., Steinbach, S., Sun, Y., Skjøth, C. A., de Weger, L. A., Lommen, S. T., Augustinus, B., Bonini, M., Karrer, G., Šikoparija, B., Thibaudon, M., & Müller-Schärer, H. (2020). Biological weed control to relieve millions from Ambrosia allergies in Europe. Nature Communications, 11(1), 1-7.

Seebens, H., Blackburn, T. M., Dyer, E. E., Genovesi, P., Hulme, P. E., Jeschke, J. M., Pagad, S., Pyšek, P., Winter, M., Arianoutsou, M., Bacher, S., Blasius, B., Brundu, G., Capinha, C., Celesti-Grapow, L., Dawson, W., Dullinger, S., Fuentes, N., Jäger, H., Kartesz, J., Kenis, M., Kreft, H., Kühn, I., Lenzner, B., Liebhold, A., Mosena, A., Moser, D., Nishhino, M., Pearman, D., Pergl, J., Rabitsch, W., Rojas-Sandoval, J., Roques, A., Rorke, S., Rossinelli, S., Roy, H. E., Scalera, R., Schindler, S., Štajerová, K., Tokarska-Guzik, B., Van Kleunen, M., Walker, K., Weigelt, P., Yamanaka, T., & Essl, F. (2017). No saturation in the accumulation of alien species worldwide. Nature communications, 8(1), 1-9.

Seebens, H., Bacher, S., Blackburn, T. M., Capinha, C., Dawson, W., Dullinger, S., Lenzner, B., Liebhold, A., Pattison, Z., Pergl, J., Pyšek, P., Winter, M., & Essl, F. (2021). Projecting the continental accumulation of alien species through to 2050. Global Change Biology, 27(5), 970-982.

Shackleton, R. T., Shackleton, C. M., & Kull, C. A. (2019). The role of invasive alien species in shaping local livelihoods and human well-being: A review. Journal of environmental management, 229, 145-157.

Shanmuganathan, T., Pallister, J., Doody, S., McCallum, H., Robinson, T., Sheppard, A., Hardy, C., Halliday, D., Venables, D., Voysey, R., Tanja, S., Hinds, L., & Hyatt, A. (2010). Biological control of the cane toad in Australia: a review. Animal Conservation, 13, 16-23.

Smart, A. S., Tingley, R., & Phillips, B. L. (2020). Estimating the benefit of quarantine: eradicating invasive cane toads from islands. NeoBiota, 60, 117.

Smith, B. J., Cherkiss, M. S., Hart, K. M., Rochford, M. R., Selby, T. H., Snow, R. W., & Mazzotti, F. J. (2016). Betrayal: radio-tagged Burmese pythons reveal locations of conspecifics in Everglades National Park. Biological Invasions, 18(11), 3239-3250.

Taylor, R., & Edwards, G. (2005). A review of the impact and control of cane toads in Australia with recommendations for future research and management approaches. A report to the Vertebrate Pests Committee from the National Cane Toad Taskforce.

Valdez, J. W. (2021). Using Google Trends to Determine Current, Past, and Future Trends in the Reptile Pet Trade. Animals, 11(3), 676.

Van Wilgen, N. J., Gillespie, M. S., Richardson, D. M., & Measey, J. (2018). A taxonomically and geographically constrained information base limits non-native reptile and amphibian risk assessment: a systematic review. PeerJ, 6, e5850.

Walsh, J. R., Carpenter, S. R., & Vander Zanden, M. J. (2016). Invasive species triggers a massive loss of ecosystem services through a trophic cascade. Proceedings of the National Academy of Sciences, 113(15), 4081-4085.

Wasserman, R. J., Dick, J. T., Welch, R. J., Dalu, T., & Magellan, K. (2019). Site and species selection for religious release of non‐native fauna. Conservation Biology, 33(4), 969-971.

No competing interests reported.

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Global economic costs of herpetofauna invasions

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Cost descriptors

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Economic costs by order and family

Cost among classes, reliability and implementation type

Cost among species and taxonomic group

Costs among activity sectors and types

Economic cost across geographic regions

Temporal trends

Discussion

Conclusion

Declarations

Author contributions

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