Risk analysis plays a crucial role in regulating and managing alien and invasive species but can be time-consuming and costly. Alternatively, combining invasion and impact history with species distribution models offers a cost-effective and time-efficient approach to assess invasion risk and identify species for which a comprehensive risk analysis should take precedence. We conducted such an assessment for six traded alien fern species, determining their invasion risk in countries where they are traded. Four of the species (Dicksonia antarctica, Dryopteris erythrosora, Lygodium japonicum, and Phlebodium aureum) showed limited global distributions, while Adiantum raddianum and Sphaeropteris cooperi had broader distributions. A. raddianum, however, was the only species found to pose a high invasion risk in two known trade countries – the USA and Australia – and requires a complete risk analysis to determine the appropriate regulatory responses. Dicksonia antarctica, Phlebodium aureum (for New Zealand), and Dryopteris erythrosora (for the USA) posed a medium risk of invasion due to the lack of evidence of impacts, and a complete risk analysis is thus deemed less crucial for these species in these countries. For other species, suitable environments were not predicted in the countries where they are traded, thus the risk of invasion is low, and a complete risk analysis is not required. For species in countries where suitable environments are predicted but no trade information or presence data are available, risk assessments are recommended to better determine the risk posed. Despite the relatively limited potential global distribution of the studied ferns relative to other major plant invaders (e.g., Pinus spp. and Acacia spp.), their history of invasion, documented impacts in pristine environments, and high propagule pressure from trade warrants concern, possibly necessitating legislative and regulatory measures in environmentally suitable regions.

early detection, environmental suitability, horticultural trade, invasion risk, MaxEnt, risk analysis

The intentional or unintentional dispersal of species into areas outside of their native range is facilitated by various pathways of introduction which may be natural or human-mediated (Wilson et al. 2009). In particular, horticultural trade has been shown to be the primary source of introductions of invasive alien plants across the globe (Dehnen-Schmutz et al. 2007a; Dehnen-Schmutz and Touza 2008; Pyšek et al. 2011; Seebens et al. 2015; van Kleunen et al. 2018). Terrestrial true ferns (Polypodiophyta; hereafter ‘ferns’) have been a popular group in horticultural trade for centuries (Birkenhead 1897; Allen 1969; Whittingham 2010) and remain popular in trade today (de Winter and Amoroso 2003). Furthermore, various aspects of horticultural trade, such as a high market presence and trade via e-commerce, directly influence the invasion success of traded alien ferns (McCulloch-Jones et al. 2021). Many horticulturally popular ferns are successfully invading and negatively impacting various parts of the globe, for example, Sphaeropteris cooperi which has extensively invaded Hawaii and displaced native populations of the tree fern Cibotium glaucum (Chau et al. 2013); and Lygodium japonicum which is classified as a species of European Union concern (EU Commission 2022) and identified as an A1/A2 pest, requiring regulation and quarantine, by the European and Mediterranean Plant Protection Organization (OEPP/EPPO 2022).

The most cost-effective means of curbing invasions is through preventative action prior to introduction (IUCN 2000; Timmins and Braithwaite 2003; Hulme 2012; Larson et al. 2020; Martinez et al. 2020) or through early detection systems promoting rapid action post-introduction or following escape from confinement (Matthews et al. 2017; Reaser et al. 2020). Risk analysis is a comprehensive process used to determine the level of risk posed by invasive alien species and directly informs the appropriate management and regulatory response. In many countries these analyses are legally required for regulatory decisions regarding alien species, for example, in South Africa, the outcomes of risk analysis are used to determine whether a species is regulated and how (e.g., through trade restrictions or permitting) (Kumschick et al. 2020). Risk analysis is a systematic and robust evaluation that comprises various components, which can be grouped into four stages: hazard identification, risk assessment, risk management, and risk communication (Kumschick et al. 2020). However, as risk analysis is expensive and labour intensive, identifying species that should undergo this process is essential to best direct scare resources. Risk assessments provide a time and cost-effective means to aid in the detection and anticipation of potentially problematic species that require attention and for which a complete risk analysis may be needed to inform regulation and legislation (Chai et al. 2016; Carboneras et al. 2018, Kumschick et al. 2020).

Risk assessments comprise the initial steps of risk analysis and generally consider the likelihood of invasion alongside consequence (negative environmental or socio-economic impacts) (Kumschick et al. 2020). Although all risk assessments generally cover these aspects, they may comprise different components depending on the approach adopted (sensu Kumschick et al. 2020). Risk assessment approaches include: trait scoring (species traits are used to predict potential for harm), statistical approaches (a trait approach combined with statistical or machine learning to determine likelihood of invasion), decision trees (a tracking system of questions and answers designed to lead to a decision), rapid screening (a quick assessment performed on a large number of species for which limited information is available), detailed approaches (a targeted approach which most closely resembles a risk analysis, requiring a substantial amount of data, management considerations, and stakeholder perceptions), or mechanistic approaches (completed for smaller groups of species evaluating the likelihood of species surpassing invasion barriers, and determining their potential impact). Recent studies applying a mechanistic approach have used similar sets of criteria to assess the risk posed by alien species, namely, a history of invasion elsewhere, invasion status (i.e., introduced and not yet naturalised, naturalised, or invasive) in the region of interest, evidence of environmental or socio-economic impact in the invaded range, and climate or environmental suitability of the region of interest (Carboneras et al. 2018; Bayón and Vilà 2019; Reaser et al. 2020).

The consideration of the climatic or environmental suitability of the receiving region for the taxon in question can greatly enhance the predictive capacity of risk assessments (Beaumont et al. 2014; Chai et al. 2016; Matthews et al. 2017; Roy et al. 2018), and can be used to inform both preventative and reactive management responses. For example, species distribution models or climate matching techniques can detect habitats or regions that are suitable for a species, but where it does not yet occur or has not yet invaded – where preventative measures can be put in place – but can also help direct detection efforts to identify previously undetected populations in suitable habitats where reactive measures may be necessary.

Species distribution modelling (SDM) is increasingly used to predict the potential distributions of alien species and identify sites that are climatically or environmentally suitable for them. These models have been extensively applied across various taxa from marine life to insects and terrestrial plants, and at various scales from local to global (Robinson et al. 2011; Kumar et al. 2015; Mainali et al. 2015; Santamarina et al. 2019; Lamelas-López et al. 2020; Venter et al. 2021). For alien ferns, SDMs have been adopted in only a small number of studies, most of which are at a fine spatial scale (i.e., at a country level or for regions within countries) (Goolsby 2004; Volin et al. 2004; Bystriakova et al. 2014; Akomolafe and Rahmad 2019). These studies have shown that alien ferns can expand their invaded ranges under present-day environmental conditions. For example, the distribution of an emergent invasive fern, Cyclosorus afer (H. Christ) Ching, is predicted to expand to an area of more than 25 000 km² across four states in Nigeria (Akomolafe and Rahmad 2018), and in Florida, the invasive fern, Lygodium microphyllum (Cav.) R.Br., could become widely established throughout the Everglades (Goolsby 2004; Volin et al. 2004).

Alien ferns are generally understudied and thus are poorly represented in official plant species inventories and in invasive alien plant regulatory lists. Therefore, to detect alien fern species that have been introduced through trade, McCulloch-Jones et al. (2023) scanned horticultural catalogues for selected trading countries (Canada, the United States of America, the United Kingdom and the Republic of Ireland, South Africa, Australia, and New Zealand). They identified a total of 382 traded fern species and determined their invasion status in the countries in which they are traded, and success or failure of invasion elsewhere. This resulted in a list of 35 alien fern species that have a history of invasion elsewhere and that are introduced but not yet naturalised in the country in which they are traded. It was concluded that these species require SDMs to better assess their invasion risk and to determine the need for a complete risk analysis in countries where they are traded.

We selected six of these alien fern species and used SDMs to determine their potential global distribution. We subsequently considered the results of these models alongside information on the species’ i) invasion status in the countries in which they are traded, ii) invasion history elsewhere, and iii) environmental or socio-economic impacts in their invaded range, to categorise each species, per trading country, in terms of the level of risk posed, and suggest the necessary response in terms of the need for risk analysis. We also indicate additional countries across the globe where the species are not yet known to occur and where risk assessment is necessary.

A broad environmental and climatic tolerance is exhibited in many popular horticultural species (Dehnen-Schmutz et al. 2007b; Hulme 2011) and thus is an attribute that is also known to promote invasion success (Hulme 2015; van Kleunen et al. 2018). As such, we expected the popular ornamental alien ferns assessed in this study to have wide environmental tolerances and, consequently, wide potential global distributions. However, the predicted distributions for most of these species are relatively limited (spanning few continents and countries), with only A. raddianum, S. cooperi, and L. japonicum showing wider potential global distributions (as compared to the other study species) both in regions where they are already known to occur and beyond their current known range of occurrence. It is unlikely that the relatively limited distributions predicted for most species are due to issues with model performance as the evaluation statistics indicated that the models performed well. Rather, this may be attributed to the habitat specificity commonly associated with fern species (Mehltreter 2008; Kessler 2010), which aligns with the suggestion that alien ferns are unlikely to invade to the extent noted for some prominent angiosperm invaders (Jones et al. 2020).

Despite the satisfactory performance of the models, a few of the response curves were truncated and slightly abnormal, indicating that the occurrence records of some of the assessed species might not characterise their full fundamental niche (i.e., the full set of conditions in which a species can survive in the absence of biotic interactions; Peterson et al. 2011) (Rodríguez et al. 2019; Soberón and Peterson 2020). This, however, is common, as the realised niche (i.e., the set of conditions in which the species occurs in the presence of biotic interactions, Peterson et al. 2011) often forms a subset of the fundamental niche; or because the occurrence records do not span the entire geographical range of the taxon (i.e., they incompletely characterise the realised niche). The latter circumstance, however, was avoided in this study as we compared the downloaded occurrence records with currently documented information on the range of each species to ensure, as far as is possible, that all areas occupied were accounted for in the models.

Variables that contributed highly to the models reflected well the documented biological and environmental limitations typical for most fern species. The large contribution of precipitation variables, for example, was unsurprising as ferns generally require moist environments (Ferrer-Castán and Vetaas 2005; Karst et al. 2005; Kessler 2010; Sharpe and Mehltreter 2010), and as precipitation plays a role in determining fern species richness (Lehmann et al. 2002; Bickford and Laffan 2006; Moreno Saiz and Lobo 2008; Nagalingum et al. 2015). The reliance of ferns on water is considered a limitation in their evolution (Page 2002), and most likely explains why terrestrial ferns are deemed unlikely to invade to the same extent as some of the world’s most aggressive invaders (see Lowe et al. 2000; Jones et al. 2019). Ferns exhibit poorly controlled evapotranspiration which restricts the degree of exposure they can withstand and limits them to areas of high precipitation and humidity (Page 2002; Sharpe et al. 2010). Accordingly, <10 % of fern species in their sporophyte form (the distinct frond producing form) exhibit desiccation tolerance (Hietz 2010). Ferns rely on water to breed as sexual reproduction can only take place where free water is available to allow movement of the motile spermatozoids involved in fertilisation (Page 2002). In accordance with these requirements, precipitation of the driest quarter was, on average, an important predictor across species. Variables associated with temperature extremes also contributed to the models for some species, and in accordance with this, the majority of fern species cannot withstand frond freezing (Fernández-Marín et al. 2021). Mean temperature of the coldest quarter, for instance, was the most important predictor for A. raddianum and this tropical species generally selects for sheltered environments where frost is unlikely. In some instances, non-climatic variables (i.e., land cover) contributed most to the models and may be an indication of habitat specificity. For example, land cover was the most important predictor for D. antarctica and P. aureum, and these species displayed highly limited potential distributions in the models. Furthermore, in South Africa, D. antarctica has naturalised in the Cape Peninsula (Roux 2001; Crouch et al. 2011), but invasion has not yet occurred (Jones et al. 2020). This is suspected to be due to its slow growth rate and late age of reproductive maturity (~20 years) (Crouch et al. 2011), coupled with fragmentation of suitable forest habitat in the Western Cape (Roux 2001).

Although many ferns can colonise disturbed and altered habitats (Murakami et al. 2005; Walker and Sharpe 2010), typical fern habitat is characterised by wet, shady, undisturbed forests (Kessler 2010). For example, in their native range A. raddianum commonly favours shaded earth banks in forest (Crouch et al. 2011; Jones et al. 2020); P. aureum occurs in forest canopies, but also in the lower strata of the forest (Smith 1993); and D. antarctica is one of the dominant understorey species of wet evergreen forests (Ough and Murphy 1996). Similarly in their invaded range, alien ferns have generally been found to largely occur in sites associated with other invaders, close to water, and most often in forest systems or in an urban/forest matrix, for example in South Africa (Jones et al. 2020), Hawaii (Chau et al. 2013), and India (Morajkar et al. 2015). Although intact forests generally display some resistance to invasion (Fine 2002; Green et al. 2004; Levine et al. 2004), various shade-tolerant and often late successional species do show the capability to invade these environments (Martin et al. 2009). Ferns are a prime example of such species (Kessler 2010; Sharpe and Mehltreter 2010): in just over six years L. microphyllum (a climbing fern) formed dense mats covering large swathes of remote intact forests in Florida, USA (Volin et al. 2004). Similarly, the related climbing fern L. japonicum (modelled here) displays comparable behaviour where it invades in northern Florida (Lott et al. 2003). Some of the other alien ferns considered in this study have been shown to outcompete native species and change forest community structure in their invaded range. For example, D. antarctica has invaded two European Union conservation priority habitats on São Miguel Island where eradication is now deemed impossible (Arosa et al. 2012), and A. raddianum has replaced populations of the native fern Adiantum capillus-veneris L. in Hawaii, USA (Wilson 1996). It is their selection for, and evidence of impacts in some unaltered environments, such as forests, that warrants concern despite their relatively limited potential global extent for invasion.

The consideration of environmental suitability in conjunction with the invasion status of a species and the knowledge of their invasion and impact history has enabled us to classify the study species in terms of the invasion risk they pose in the countries in which they are traded, and thus get an indication of whether they require a complete risk analysis. The USA, Australia, and New Zealand show the greatest potential for invasions by these traded alien ferns, with suitable environments available for at least four of the six study species which now constitute medium or high risk species that require risk analysis. The USA is of particular concern, as this country already has extensive documentations of fern invasions and is a prominent trader in alien fern species (McCulloch-Jones et al. 2023). For example, in the USA, the Florida Exotic Pest Plant Council’s 2019 List of Invasive Plant Species lists seven fern species as invaders that are changing community structures or ecological functions, or hybridizing with native species (FLEPPC 2019). Furthermore, in Georgia, USA, five alien ferns have recently been documented as spreading beyond their sites of introduction (Wyatt 2020), including D. erythrosora (modelled here) which was first noted outside of cultivation just 13 years ago (Simpson et al. 2008). According to our study, Georgia and its surrounding areas are environmentally suitable for this species, but the area available for range expansion is closely confined to its current invaded range (which spans a large area of south-eastern USA, excluding Florida). Similarly, New Zealand also has a history of multiple fern invasions (Brownsey 1980; Brownsey and Perrie 2014; Perrie et al. 2019) and there are suitable environments available in New Zealand for three of the study species, namely D. antarctica, L. japonicum, and S. cooperi, but these have a relatively limited potential range, and moderately suitable environments.

The USA, Australia, and New Zealand are most in need of trade regulations, specifically for the species identified as medium and high risk to prevent the occurrence of, or increase in, invasions. As official regulation is contingent on risk analysis, this study provides crucial information by identifying species that require risk analysis. Efforts to detect and manage escaped populations are necessary as suitable environments are available and invasive populations may remain undetected in these countries. Furthermore, countries that possess suitable environments, but lack documented occurrences of these species should conduct risk assessments. Based on the assessment outcomes, they can determine whether it is necessary to perform risk analysis and implement trade regulations to prevent future invasions. It is important to note that all species considered in this study are highly popular in trade and introduction via this pathway is thus likely (McCulloch-Jones et al. 2021). We suggest that key habitats (e.g., moist, shaded environments) that occur close to, or as a matrix with, urban areas are prioritised for efforts to locate undetected populations as this is where fern invasions most commonly occur. Trading countries for which the majority of species were categorised as a low risk (i.e., Canada and the United Kingdom and the Republic of Ireland) are under a lower potential threat of invasion and the current response arguably need not be as aggressive (i.e., risk analyses are not indicated, and trade bans are likely unnecessary). Importantly, the suggested categorisation of the study species should be revised periodically, given changes in land-use and climate, and as the occurrence and environmental data available for modelling could improve with time as further resources become available.

This study is the first to model the potential global distribution of multiple invasive alien fern species, identifying countries susceptible to invasion and informing the need for risk analyses. The models reveal relatively limited potential global distributions for these ferns compared to other major invaders, but their association with undisturbed habitats, such as forests, and their documented impacts in such habitats raises concern. The projections of environmental suitability have allowed us to complete risk assessments for the studied species to inform their requirement for risk analyses in the countries in which they are traded, thus initiating the early stages of management action. The USA, Australia, and New Zealand stand out as potential hotspots for invasion by traded alien ferns, necessitating management interventions and on-ground population detection for high or medium risk species. This risk assessment approach serves as a valuable management tool, highlighting focal species for each trade country and supporting efficient resource allocation in alien species management and regulation.

We thank all data collection assistants for their time and help with the study. Support and funding for this work was provided through the DSI-NRF Centre for Excellence for Invasion Biology (C•I•B) and the South African National Biodiversity Institute (SANBI). Katelyn T. Faulkner and Emily J. McCulloch-Jones thank the South African Department of Forestry, Fisheries and the Environment (DFFE) for funding, noting that this publication does not necessarily represent the views or opinions of DFFE or its employees.