Black, Grey and Watch Lists of alien species in the Czech Republic based on environmental impacts and management strategy

As legislation, research and management of invasive alien species (IAS) are not fully coordinated across countries or different stakeholder groups, one approach leading to more or less standardized activities is based on producing lists of prominent IAS that attain high level of concern and are a subject of priority monitoring and management. These so-called Black, Grey and Watch (alert) Lists represent a convenient starting point for setting priorities in prevention, early warning and management systems. It is important that these lists be based on transparent and robust criteria so as to accommodate interests and perception of impacts by groups of concerned authorities and stakeholders representing sectors as diverse as, e.g. forestry, horticulture, aquaculture, hunting, and nature conservation, and to justify possible trade Copyright Jan Pergl et al. This is an open access article distributed under the terms of the Creative Commons Attribution License (CC BY 4.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited. NeoBiota 28: 1–37 (2016) doi: 10.3897/neobiota.28.4824 http://neobiota.pensoft.net RESEARCH ARTICLE Advancing research on alien species and biological invasions A peer-reviewed open-access journal NeoBiota Jan Pergl et al. / NeoBiota 28: 1–37 (2016) 2 restrictions. The principles for blacklisting need to be general enough to accommodate differences among taxonomic groups (plants, invertebrates, vertebrates) and invaded environments (e.g. aquatic, terrestrial, urban, suburban, seminatural), and must take into account invasion dynamics, the impact the IAS pose, and management strategies suitable for each particular invader. With these assumptions in mind, we synthesize available information to present Black, Grey and Watch Lists of alien species for the Czech Republic, with recommended categorized management measures for land managers, policy makers and other stakeholders. We took into account differences in the listed species’ distribution, invasion status, known or estimated environmental impact, as well as possible management options, and apply these criteria to both plants and animals. Species with lower impact, but for which some level of management and regulation is desirable, are included on the Grey List. Some potentially dangerous species occurring in European countries with comparable climatic conditions, as well as those introduced in the past but without presently known wild populations in the Czech Republic, are listed on the Watch list. In total, there are 78 plant and 39 animal species on the Black List, 47 and 16 on the Grey List, and 25 and 27, respectively, on the Watch List. The multilayered approach to the classification of alien species, combining their impacts, population status and relevant management, can serve as a model for other countries that are in process of developing their Black Lists.


Impacts of invasive alien species and Black Lists: state of the art
Although only a small proportion of introduced species become naturalized or invasive and have a measurable impact (Lockwood et al. 2013; but see Ricciardi et al. 2013), biological invasions by alien species (introduced to regions outside their native distribution range due to human activities; Richardson et al. 2000, Blackburn et al. 2011) affect the majority of habitats, including semi-natural ones. Invasive alien species (IAS), with their widely documented impacts on biodiversity, ecosystem functioning and economy (Pyšek and Richardson 2010, Vilà et al. 2010, Pyšek et al. 2012c, Scalera et al. 2012, Follak et al. 2013) are recognized as one of the key components of global environmental change (MEA 2005). Costs due to IAS were estimated to reach up to 5% of global GDP (Pimentel et al. 2001(Pimentel et al. , 2002. In Europe, recent estimates of direct costs due to IAS reach at least 12.7 billion € per year ). It is also important to note that direct environmental and eradication costs associated with environmental weeds or pests are only a small fraction of costs caused to agriculture or forestry. Nevertheless, even these figures on overall costs for environmental weeds and pests illustrate the need for an urgent policy response at all scales, from national to international and global, supported by a corresponding scientific knowledge base; the fact that the majority of alien species are introduced intentionally or in association with imported/transported commodities (Hulme et al. 2008(Hulme et al. , 2009) provides an opportunity for interventions (Roques and Auger-Rozenberg 2006, Dehnen-Schmutz et al. 2007, Kenis et al. 2007).
In Europe, more than 12,000 alien plant and animal species are recorded (DAISIE 2009, www.europe-aliens.org) and the numbers of successfully establishing species continue to grow (Hulme et al. 2009, van Kleunen et al. 2015. Unfortunately, research, legislation, and management of IAS are not fully coordinated, neither within individual countries, nor continentally (Hulme et al. 2009), which leads individual countries to cope with alien species in different ways. The most common approach is based on producing lists of prominent IAS that receive much attention and are prioritized in terms of prevention, monitoring and management. These so-called Black, Grey and Watch (alert) Lists represent a convenient starting point for setting such priorities (European Commission 2014). The necessary condition for making such lists trustworthy is, however, a robust and transparent risk assessment, based on the impacts of individual species, allowing their scientifically defensible selection (Wittenberg and Cock 2001, Verbrugge et al. 2012, Lewis and Porter 2014. The transparency is important so as to accommodate interests and perception of impacts by groups of concerned authorities and stakeholders representing sectors as diverse as, e.g. forestry, horticulture, aquaculture, hunting and nature conservation, and to justify possible trade restrictions (Bayliss et al. 2013, Kelly et al. 2013, Ööpik et al. 2013. Therefore, when developing regional Black Lists, interests that differ among the above-mentioned sectors need to be taken into account. Many intentionally imported alien species are of a high economic value (DiTomaso et al. 2010, Richardson and Rejmánek 2011, Woziwoda et al. 2014), but can have negative impacts on native populations, species and communities due to a wide range of mechanisms and processes that have been described in the literature in the last decade (e.g. Levine et al. 2003, Gaertner et al. 2009, Mitchell et al. 2010, Pyšek and Richardson 2010, Vilà et al. 2010, Dodet and Collet 2012, Pyšek et al. 2012c, Scalera et al. 2012. However, although these processes are becoming reasonably well understood, there is still much uncertainty about which particular species will have an impact in specific environmental settings and how the invaded habitats and ecosystems will be impacted (Leung et al. 2012. Ideally, each intentional introduction of a new alien species should be thus preceded by a cost-benefit analysis of negative vs. positive effects on both the environment and socioeconomy (Keller and Drake 2009). The decision should then reflect the climatic and habitat match between the current range of the species and the region to which it is proposed for import, as well as information about previous invasion history and life history traits of the species itself, or its close relatives Lodge 2001, Keller andSpringborn 2014).

IAS regulation in Europe and in the Czech Republic
The urgent need to tackle biological invasions, develop a common policy and establish an early warning system in Europe, has been recognized by the European Commission (see the Communication 'Towards an EU Strategy on Invasive Species', (COM (2008) 789 final) and EU Biodiversity Strategy to 2020 (http://ec.europa.eu/environment/ nature/biodiversity/comm2006/2020.htm). Part of this activity is aimed at the new EU Regulation on IAS COM (2013) 620 (European Commission 2014), which is an important legislation on invasive species threatening biodiversity and human well-being (Genovesi et al. 2015). Besides setting a framework for roles and responsibilities among the different bodies dealing with IAS it will include a list of species that pose the most significant threats (list of alien species of the Union concern) and thus should be prohibited from the import, sale, and use in Europe. This list will be prepared by the European Commission on the basis of the criteria set out in the Regulation; the EU member states participate in the process of the preparation of the list (by providing comments and proposals for individual IAS inclusion). Although national Black Lists may play an important role in the process of preparation of the EU list, so far only a few countries have developed their own Black Lists with some legislative support .
The development of national and regional Black Lists and identification of important species, based on using standard and transparent criteria, is a key aspect of the early warning and information systems. Some European countries or trade sectors (agriculture, aquaculture) already regulate the introduction and transport of selected species, based on risk assessments provided by the European Plant Protection Organisation (EPPO), European Food Safety Authority (EFSA) and UK Department for Environment, Food and Rural Affairs (DEFRA). An example of a working system is international cooperation in the field of agriculture pests (EPPO, DEFRA) which can serve as a template to be followed for the management of IAS in Europe in general (Brunel et al. 2013). Not only legislative tools are affecting the policy on IAS. To prevent the spread of alien species and restrict their trading, a significant component of policy and public involvement are voluntary codes of conduct developed for example for horticulture or sheltered under the Bern Convention (Heywood and Brunel 2011, Caffrey et al. 2014, Halford et al. 2014, Heywood 2014.
In the Czech Republic (78,866 km², 10.5 millions of inhabitants), as in many other European countries, there is an elaborate and legislatively well-anchored system of the approach to harmful organisms in agriculture. In the field of nature conservation, legislation is not sufficient and does not adequately respond to the current threats from biological invasions, but the issue of IAS has become in the last years one of the priorities in the Czech national strategic environmental documents (State Environmental Policy 2012-20, State Programme of Nature and Landscape Conservation 2009, Biodiversity Strategy 2005. These documents emphasize the need to focus on IAS, including development of priority lists of species for management, creating financial tools and preparation of new legislation, which will be encouraged by the adoption of the new IAS EU legislation.

Scoring species for Black Lists
Despite significant progress in producing lists of important alien species for individual countries (see review in Essl et al. 2011), a standard methodology for the complex as-sessment of their impacts only started to appear recently (e.g. Blackburn et al. 2014). Such a framework needs to be accompanied by a close cooperation between policy makers, researchers and practitioners in nature/biodiversity conservation and IAS management, to allow for harmonization of the information flow on IAS (Ricciardi et al. 2000, Caffrey et al. 2014. Species with documented strong negative impacts, that threaten ecosystems, habitats or native biota, should be eradicated from the newly invaded sites as fast as possible, and further introductions of such species avoided (Convention on Biological Diversity 1992, Genovesi 2005. However, if resources are limited, the question remains which species, which locations and how (considering feasibility and control methods) should be targeted first, and this prioritization can be addressed by different methods (Humair et al. 2014).
The criteria for placing individual species into particular Black List categories need to be general enough to accommodate differences among various taxonomic groups (plants, invertebrates, vertebrates) and invaded environments (e.g. terrestrial, aquatic; urban, suburban, seminatural), take into account invasion dynamics, the environmental and socio-economic impact they pose and management strategy suitable for each particular invader. The existing Black Lists do not take differences between invaded habitats and management feasibility into account in their assessment, do not cover socio-economic impacts and are restricted to selected taxonomic groups . Some of the existing impact assessments, serving as a basis for Black Lists, multiply the impact scores by a given species' population status (Gederaas et al. 2012, http://ias.biodiversity.be) but as far as we know, there is no system that incorporates information on the type of invaded habitat and management feasibility into the Black List classification.

Aims of the study
In the Czech Republic, there is a thorough knowledge of biological invasions that has resulted in publications of comprehensive and updated lists of alien plants and animals (Pyšek et al. 2002, Šefrová and Laštůvka 2005 with an indication of their invasion status using commonly accepted classification (Richardson et al. 2000, Pyšek et al. 2004, Blackburn et al. 2011. However, the classification of alien species based on management criteria has not been available up to now. Still, for any management planning, setting the priorities among species and habitats is crucial. In this paper we thus combine information on the potential environmental impact of alien species in the Czech Republic, their current or predicted population status, the feasibility of management, and type of invaded habitats. As a synthesis, we present Black, Grey and Watch Lists of alien species for the country, with recommended categorized management measures for land managers, policy makers and other stakeholders.

Data sources and species selection
The proposed Black and Grey Lists of alien species in the Czech Republic are based primarily on the existing inventories of plant ) and animal (Šefrová and Laštůvka 2005) alien species. The data from these lists were amended by recent updates of the alien biota in the Czech Republic for particular groups such as fishes (Musil et al. 2010), national museum collections or unpublished records (personal communications and databases). The Watch List of alien species includes those currently not present in the wild in the Czech Republic and occurring there only in captivity or cultivation, but reported from the wild in other European countries with similar climate and habitats. Existing lists of aliens in these comparable countries, as summarized in e.g DAISIE or Nobanis, were thus screened to generate the Watch List for the Czech Republic.
To minimize the possible subjective bias of experts assessing species on original lists, each species was reassessed according to the current state of its population status, invaded habitats, cultivation and farming history, impact on environment (ecology) and socio-economy and with respect to the knowledge of its effective management. The species sharing similar patterns of classification were then grouped into subgroups of Black and Grey Lists (see details below). Species included in Black Lists were those posing significant strong negative effects on the environment and where some management, if available and feasible, should be applied. Grey List was used for species with limited negative environmental impact, where monitoring and local management is also relevant. Species for Watch List were selected from those that may in the near future colonize the territory Czech Republic and whose monitoring and management, due to possible substantial negative environmental impact, is recommended.
The evaluation of alien species occurring in the Czech Republic was done for vascular plants, vertebrates and most invertebrate groups. As the classification of alien plant species in the Czech Republic is more elaborated than that of animals, in terms of their regional population dynamics or abundances (Pyšek et al. 2012a, b), the criteria for the Black List species' assessment were first developed for plants and then adapted for other taxonomic groups.

Criteria for classification
For each species included in the Black, Grey and Watch List based on the above criteria, the following information on their populations was assessed, if available, and used to classify species.
A. Mode of current spread: 1. Plants and animals that are intentionally released into the environment for landscaping, restoration or hunting (the 'release' pathway according to Hulme et al. 2008) and distribution of the species is highly dependent on human activities. Without presence of humans activities the species will disappear in relatively short time. 2. Current spread is mostly spontaneous without direct contribution of humans. For this category it is not crucial if the initial occurrences resulted from past human activities (abandoned plantations, populations of animals escaped from cultures, contaminants) or results of spontaneous spread from other areas where they are alien. Without presence of human activities the species will remain in the landscape for relatively long time. 3. Combination of release and spontaneous spread.

B. Distribution:
Current distribution regardless of whether the species occurs as a result of release or spontaneous introduction. This categorization does not take into account abundance of the species. Both groups can be represented by dense or sparse populations. Especially in case of regionally widespread species, which are present in numerous, well established and continuously replenished populations, their local management cannot be usually efficient. However, in some cases local management may still be performed to reduce specific impacts, e.g. local and time-restricted trapping of Neovison vison (American mink) before the bird breeding season. 1. Regional: Present distribution of the species at a large scale or future expasion not strongly restricted by environmental constraints is expected. Clusters of local populations dispersed across country exchanging individuals due to the transport of propagules or active migration. 2. Local (isolated populations): current and also future distribution in localized area(s) within the Czech Republic. The distribution can be limited by e.g. climate or habitat specificity. The localized distribution makes management efficient if there are effective methods available.

C. Evaluation of environmental impact
Standardized assessment of environmental and socio-economic impact is not available for all alien species in the Czech Republic. Therefore it was assessed using the simplified rationale of GISS (Nentwig et al. 2010, Kumschick et al. 2012, Vaes-Petignat and Nentwig 2014 and the recently suggested unified classification of alien species based on the magnitude of their impacts ). The black listing in this study is based primarily on the environmental impact of populations occurring in the outdoor environment, and excludes e.g. alien species only having significant economic impact as storage pests. Due to the lack of direct knowledge on impacts of many species in the Czech Republic, their impact was classified as "potential impact", taking into account any impact of the given species reported from climatically similar regions, and also considering interactions with, or impact of, ecologically similar species. The impact was classified based on expert judgement into three levels ranging from limited (minimal) to moderate and massive, with respect to whether it results in irreversible negative changes to native populations, species or ecosystems (e.g. due to predation, competition, hybridization, ecosystem functioning). For impact assessment we used data from Kumschick et al. (2015), and Rumlerová et al. (unpublished).

D. Evaluation of socio-economic impact
Socio-economic impact and impact on humans was additionally assessed for taxa with considerable environmental impacts to support final reasoning of recommended management. The weight of socio-economic impact was used and ranked high in case of species like Ambrosia artemisiifolia (common ragweed), Heracleum mantegazzianum (giant hogweed), where strong negative impact on human health is significant or Arion vulgaris (Lusitanian slug), and Varroa destructor (varroa mite), which have direct effect on agriculture. The impact was classified based on expert judgement into three levels ranging from minimal to moderate (most weeds and pests) and massive.

E. Management options
Management options were assessed along axes representing the management itself, the context of invaded habitats, and population status. The species were classified according to the applicable management strategy (see details below and in Table 1).
Complete eradication is hardly feasible in the Czech Republic, an inland state surrounded by other countries, and can be only achieved, if at all, by intensive international cooperation followed by continuous sanitary measurements. Although complete eradication is usually feasible only on islands (e.g. Chapuis et al. 2004, Genovesi 2005, Simberloff et al. 2011, in some cases it is an ideal target to which efforts should be directed. In practice, complete eradication is possible only for populations of alien species that do not yet spread. For large infestations consisting of many metapopulations, complete eradication above some threshold is almost impossible due to enormous costs Pitcairn 2002, Pluess et al. 2012a, b). High cost of management can be justified only for newly detected occurrences of highly important alien species. Unfortunately, intentions behind eradication attempts are often led by wrong ideas to restore ecosystems to their "historical" state, which is often idealized. Eradication is sometimes initiated by the local public or little-informed conservation activists, and often is accompanied by damages to native communities.
Tolerance (resignation) means to refrain from any systematic attempts to manage the given alien species; although both lead to the same result, reasons for them are fundamentally different: tolerance is result of a decision based on the fact that the given IAS has a low impact, while resignation is an enforced attitude if there are no existing management options. The latter currently happens in e.g. mine disposal sites in northern Bohemia, where management is passive approach, and eradication efforts focused on a few selected plant species and habitats. Many newly introduced plants continuously spread as a result of restoration of brownfields and landscaping (Kabrna et al. 2014). Similarly, for some insects, e.g. Harmonia axyridis (harlequin ladybird), any management action is almost impossible.

Tolerance/ resignation
This approach is relevant in many ecosystems/sectors (forestry, fishery) for several reasons. Many alien species occurring now in the landscape are of a high economic importance. This approach is also relevant for large populations of widespread alien species especially in urban and suburban environments. Direct eradication of such species is almost impossible or associated with enormous costs and likely to bring doubtful results.
Tolerance is applicable in several cases. In some urban and suburban areas we recommend to tolerate the species of a high economic value as well as species eradication of which is almost impossible because of their wide distribution. This tolerance should exclude areas of high conservation value where approaches including local eradication with subsequent change of local management can be applied. Tolerance cannot be used in rural landscape where primary aim is to prevent new alien populations from establishing. We recommend to tolerate e.g. large populations formed as a result of old abandoned plantations (e.g. Robinia pseudoacacia) or release (crayfish, white-tailed deer).

Eradication
Complete eradication of alien species at national scale. It is usually demanding in terms of financial, time and human labour resources, and would require transboundary coordination in case of species present also in neighbouring countries.
Complete eradication should be used primarily for small and pioneer populations where rapid response is likely to result in successful action. It is also applicable to small populations of relatively large animals where hunting or other effective control is feasible. Eradication is not recommended in urban and suburban environment where it usually fails for several reasons (public opinion, high propagule pressure). The complete eradication of several species currently posing strong negative socio-economic impact can be reasoned.

Containment
Local eradication or suppression of alien species' populations. Depending on infested area and habitat type, the costs can vary. Repeated and continuous management is necessary to meet the goals.
Containment is recommended only for sites with high conservation priorities or to lower the negative impact of selected alien species. Due to high costs and need to repeat the actions regularly it is not recommended in large areas, or urban and suburban environment. Containment can be used to reduce e.g. the propagule pressure.

Removal of populations from abandoned plantations and farming facilities
Removal of populations after cessation of their planting or farming, especially related to biofuel plants and animals bred in cages, fishponds or forest enclosures.
Complete eradication of the populations at local scale is recommended, as there is a high risk of escape into natural environment following the abandonment.

Prevention of spread to (semi-)natural environment
This management option refers mainly to revegetation activities in suburban zones (along road and railway corridors) and to species released for forestry, game hunting or fishery.
This option should be used in most cases to avoid conflicts of nature conservation with forestry, landscaping, agriculture and hunting. If a release of a species into the wild is considered, preference should be given to native or locally native taxa. Examples are e.g. brown vs rainbow trout, or red vs sika deer.

Change of management
Change of management is a widely used method applicable to a wide range of habitats. In rural landscapes such a recommended management (preferred by nature conservation) is similar to the traditional management (regular mowing, removal of shrubs, grazing). This management option includes also hunting and fishery practices.
In case of plants, change of the current management should be used to reduce the cover and therefore impact of local dominants. Important condition is that the management has to be permanent and resulting ecosystem must be of higher natural quality than the previous one. Change of management is relevant for a wide range of stakeholders including forestry, game hunting and fishery.
At present we are unable to stop the invasion of such species, let alone eradicate them completely.
Stratified approach reflects the local/regional context of the invasions and therefore represents, in the vast majority of cases, the optimal strategy. An example is the management of Robinia pseudoacacia (black locust) in the Czech Republic, whose planting can be allowed in areas where the stands do not represent an imminent threat to the landscape, but should be prohibited, and extant stands eradicated, from sites with nature conservation needs, such as in and around steppe habitats. Similarly, some economically important alien fish species are tolerated in aquaculture ponds (many of which are localities of high conservation value, and even listed among protected nature reserves and Natura 2000 sites), but in other localities might be subject to management. For example, the native Salmo trutta (brown trout) should be preferred over alien salmonid fish, such as Oncorhynchus mykiss (rainbow trout), in stream habitats, but alien fish species are less likely to pose a conservation problem in ponds used for recreational fishing. The stratified approach thus discriminates where and when the management of alien species is needed and efficient, and where the eradication is neither effective, nor necessary (e.g. in urban and suburban areas). The stratified management limits counterproductive and useless actions against alien species and places them into the framework of nature protection and traditional land use management.

Results
Although there are differences in life histories, population status and possible management options between plants and animals, in the proposed scheme for blacklisting we were able to produce comparable Black, Grey and Watch lists for these groups together. In the Black List, species were assigned into three categories according to their impact, distribution, population dynamics and management strategy (Table 2). It is important to note that individual subgroups of Black Lists do not reflect the importance of the included species in the descending order. Species listed in the Grey List have lower impact than Black-Listed species, but still may require some level of management and regulation. The eradication of Grey-List species at a large scale is not a high priority, nevertheless their management is recommended in some restricted areas with nature protection concerns. Grey and Watch List species should be monitored for any rapid change in their distribution and possible impact, especially on the environment.
In total, there are 78 plant and 39 animal species on the Black List, 47 plant and 16 animal species on the Grey List, and 25 plant and 27 animal species on the Watch List (Appendix).

Black and Grey Lists of alien species in the Czech Republic
There are in total 1454 alien vascular plant species recorded in the Czech Republic (36.6% of the total flora; Pyšek et al. 2012a, b), however, the vast majority of them do not have a measurable impact. This group of "low impact species" consist of species that (i) are unable to reproduce or develop viable populations outside cultivation (casuals); (ii) are naturalized but have not expanded their range for a long time, or even failed to persist and became rare (e.g. Agrostemma githago, common corn-cockle) and (iii) are locally naturalized, having potentially negative impact (e.g. Celastrus orbiculatus, oriental bittersweet), but their sparse distribution still makes management feasible. Within the last group belong species which are candidates for priority monitoring (e.g. biofuel plants like Paulownia tomentosa, princess tree). Alien plant species with potentially high risk of environmental and potential negative socio-economic impact thus recruit from naturalized species starting to spread (85 species), or species with continuing spread (61 species). The assessment of fauna was based on several sources providing an overview of alien animal species occurring in the Czech Republic: 662 species from the DAISIE database (Pergl et al. 2012), 595 species from the catalogue of alien animal species (Šefrová and Laštůvka 2005), and 490 species from the list of alien terrestrial insects occurring in indoor and outdoor environments (Šefrová 2005 and unpublished database of Šefrová et al.). This screening resulted in a total of 680 alien animal species, the majority of which are terrestrial insects (490), followed by other terrestrial and aquatic invertebrates (110) and vertebrates (80). Of the alien terrestrial insects, 249 are known to be restricted to indoor spaces where stable temperature allows them to shelter from harsh winter conditions outside, and the same holds for the majority of arachnids and gastropods. These species, unable to escape into the outdoor environment, were thus not included in the assessment for the Black List. As a result, we identified 184 animal species that occur outdoors and have (or potentially may have) an environmental impact.
There are 102 established (naturalized) but not invasive insect species that have not spread significantly or had already spread in the past and now are considered as a part of resident communities. Among the invasive insects, seven species have an impact on native insects and 41 can be classified also as pests in agriculture, forestry or horticulture. Of these, 28 species cause significant losses to the economy and are therefore permanently monitored and managed; monetary value of the damage to the environment, if at all possible to estimate based on current knowledge, is by an order of magnitude lower than that to economy.
In the list, we retained two invertebrate species known to have more devastating effect on agriculture than on biodiversity, Arion vulgaris (Lusitanian slug) and Varroa destructor (varroa mite), which potentially can also have a strong environmental impact. Arion vulgaris is generally widespread and may influence also natural communities by herbivory and competition with native gastropods; the environmental impacts of V. destructor are indirect, through its potential effect on the pollination by honeybees. In aquatic environments, the proportion of invertebrates with possible impact on native species or ecosystems is relatively high, with representatives from macrozoobenthic molluscs, such as Dreissena polymorpha (zebra mussel), or crustaceans, such as the amphipod Dikerogammarus villosus (killer shrimp), or invasive crayfish (Orconectes limosus, spiny-cheek crayfish; Pacifastacus leniusculus, signal crayfish).
Alien vertebrates are the smallest group in terms of species number, but host the highest proportion of species causing ecological impacts. There are marked differences among vertebrate groups. There is no alien bird with negative ecological impact in the Czech Republic, and only one reptile (Trachemys scripta, pond slider), which so far does not seem to be able to reproduce in the wild under the local climatic conditions. In contrast, fish and mammals with well documented or potential impact are quite common. Several of these fish (~10 spp.) and mammals (~15 spp.) are already widely distributed in the Czech Republic, and their complete eradication is not feasible. However, local/regional eradication or suppression by management action may be possible. It is therefore important to reduce new introductions and releases and strictly control the vicinity of farming and breeding facilities (e.g. deer parks, fishponds) to prevent or at least diminish escapes into nature.
The groups of alien species classified within the Black (BL1-3) and Grey Lists are characterized mainly by level of impact, type of spread (affecting the management and regulation). Species with high environmental and high socio-economic impact are in BL1. Species with high or medium environmental impact and almost negligible socioeconomic impact are then classified according prevailing mode of their spread (BL2, BL3). Species, the environmental impact of which is limited at present, are included in the Grey List (GL). The detailed description of the groups is following: Species group BL1: Species with the greatest impact and with the strongest regulations recommended/needed; their populations should be managed whenever possible although they are already present in large numbers in the Czech Republic and their complete eradication is not feasible. Whenever feasible, it is important to limit further spread of these species; for species where efficient management strategy is not available at present, research that may provide management options is warranted. The group includes two plant and three animal taxa. Plants listed in these category are rapidly spreading neophytes, an annual Ambrosia artemisiifolia (common ragweed) and monocarpic perennial Heracleum mantegazzianum (giant hogweed), having strong impacts on native biodiversity and/or posing direct threats to human health (allergy and photodermatitis) (Nielsen et al. 2005, Hejda et al. 2009. Animal taxa comprise heterogeneous group of species which include Varroa destructor, a mite affecting bees, and two mammal species (Neovison vison, American mink; Procyon lotor, racoon). As Varroa has also significant socio-economic impact and is restricted to honey bee colonies, its distribution is monitored and management is already driven by state authorities.
Species group BL2: Species depending highly on human actions that promote their spread (mostly combination of release and spontaneous spread), both types of distribution, and mostly with moderate to massive environmental impact, but minimal socio-economic impact; 49 plant and 8 animal taxa. These species are often found as remnants of planting in gardens and plantations or in case of animals introduced for hunting and fishing, which facilitates their further spread. Instead of economically important species, alternative native species should be promoted. If necessary for economic activities in areas with low conservation value, keeping in capture could be permitted, with prerequisite of good prevention of escape, and removal of the captive population once the economic activity has ceased. Spontaneous populations outside urban areas or areas of captivity should be reduced by change of local management, or by local eradication campaigns when feasible. Specific focus should be on areas with high conservation value.
Species group BL3: Species whose current distribution results from spontaneous spread and unintentional introductions. They cover species with both types of distribution and impact ranging from limited to massive (Appendix). The recommended strategy for these species is stratified approach balancing between the local needs and the available resources for eradication. As none of the species is planted or released intentionally, the management and trade regulations can be more straightforward than in BL2. If locally necessary and there are known efficient eradication methods for the given species, eradication should be attempted. In urban and suburban environments species can be tolerated, but eradication or suppression by change of local management (land use) is recommended.
Species group GL: Species with limited environmental impact at present, distributed both regionally and locally, and with current distribution as a results of spontaneous or combined spread. For the listed species outside areas of a high conservation value there is no need to take actions against them, or restrict them. Change in management may be actively taken into account to reduce their distribution. This group consists of 47 plants and 16 animals, and is substantially formed of several weedy plant species and parasites.

Watch List of alien plant and animal species
The Watch List (Appendix) contains selected high-impact species that (1) have not yet been recorded from the Czech Republic but occur in other European countries with similar climatic conditions and habitats (and thus may be successfully introduced to or invade the Czech territory), (2) species that are at present kept in culture or enclosures only (such as Capra aegagrus, wild goat, or Bison bison, American bison), or (3) species introduced in the past but without presently known wild populations, which may be considered potential competitors for native species (several fish species). In case of plants this is analogous to species already present in e.g. gardens, parks or aquaculture (e.g. Azolla filiculoides, Pacific mosquitofern; Paulownia tomentosa, princess tree) which may in the future establish in the wild and became problematic. There are 25 plant and 27 animal taxa on the Watch List. For these species, as well as for some sparsely distributed species from the Black or Grey Lists, preventive actions against their introduction to and subsequent spread in the country, or uninvaded regions, are justified.

Discussion
This paper provides the first assessment of alien species in the Czech Republic in terms of their environmental impact, with direct habitat-related recommendations for land managers, policy makers and other stakeholders. Introduction and naturalization of a new species is a dynamic process (Blackburn et al. 2011, Richardson and Pyšek 2012, Lockwood et al. 2013, therefore the published lists of this kind are not and cannot be definitive. One of the important aspects of such a work is that it can stimulate discussion on the assessment of individual species as well suggestions of possible additions or deletions, from people involved in research, management, as well as general public. It has to be highlighted that the proposed groups BL2 and BL3 within the Black List do not show the importance of the included species for prioritization of the management as their environmental impacts, though not negligible, may vary. The grouping is used mainly to differentiate between various management options in respect to particular site conditions. Furthermore, these lists are based on environmental rather than socio-economic impact. Thus, we did not include in the list pests causing heavy economic losses, like Leptinotarsa decemlineata (Colorado potato beetle), the impact of which is restricted exclusively to agriculture. In contrast, we included, for example, Varroa destructor, whose impact on commercial honey bees may have indirect environmental consequences through effects on pollination of many plant species.
Within the Grey List, we included also a taxon that, despite being a part of the alien fauna in the Czech Republic, does not require management in the wild but rather import restrictions. This is the case of the Chinese mitten crab (Eriocheir sinensis), a potential host of a serious pathogen that can be transmitted to freshwater crayfish, i.e., native species of conservation relevance (Svoboda et al. 2014). Due to its transient occurrence in the Czech Republic (during periodic migrations only), this species was not listed in group BL3 that includes alien crayfish species with the same capability but established in the country and thus eligible for local management. For the Chinese mitten crab, a legislative ban of release into the wild as well as regulation of trade and import of live individuals are recommended; if an import is considered, only dead animals for food market should be imported.
The system presented here follows the recommendations of IUCN that all newly introduced alien species should be treated as "guilty until proven innocent", following the precautionary principle (Genovesi 2005). The proper evaluation of a species is hindered by a possible lag phase between the introduction and naturalization (Williamson et al. 2005, Blackburn et al. 2011) and a wide range of possible impacts that are context-dependent (Pyšek et al. 2012c, Hulme et al. 2013, Horáčková et al. 2014. In reality, the recognition of problematic invasive alien species in early stages is very difficult and usually not possible until the species is widely distributed; at that stage, however, it is usually too late for its easy eradication (Pluess et al. 2012b).
Invasive alien species are responsible for many negative effects on native species and ecosystems, particularly in areas with a high conservation status (Foxcroft et al. 2013) where IAS management is costly and makes up a large proportion of the pro-tected area management budget (Frazee et al. 2003). In contrast, in many ecosystems, human activities and resulting land-use change, such as increasing intensification of agriculture and urbanization, or abandonment of industrial areas, promotes existence of "novel" habitats where some alien species might be a valuable component (Hobbs et al. 2006, Gaertner et al. 2012. This is the case of green areas in and around cities where the native species diversity is reduced and vegetation is composed of a few dominant native species accompanied by aliens with a relative low cover. Urban areas are a significant source of alien species (Aronson et al. 2014, Kowarik et al. 2013), but they also fill important ecosystem services with wide socio-economic implications. Therefore, to eradicate or not is often not a simple decision, especially if one takes into account financial costs and feasibility of such a management action.
A separate issue related to alien species and our proposed Black, Grey and Watch Lists are recent developments in the area of biofuel plants and animal species imported for aquaculture and farming. It has been suggested that the traits of an ideal biofuel species are the same as those favouring invasiveness (Raghu et al. 2006, Buddenhagen et al. 2009, Smith et al. 2015. Some of the biofuel species (Arundo donax, giant cane; Psidium cattleianum, cattley guava) are even listed among 100 of the worst global invaders of the IUCN (Lowe et al. 2000). In the Czech Republic, the issue of importing and planting potentially invasive species is manifested by the biofuel or forestry species such as Reynoutria taxa, or Quercus rubra and Paulownia tomentosa, respectively. For such cases, we advocate a stratified approach based on the type of the invaded habitat, and habitat-related nature conservation needs. A knowledge-based and region-specific differentiated approach is much more suitable than efforts aimed at complete eradication, regardless of circumstances, which is in most cases hardly possible anyway (Rejmánek and Pitcairn 2002, Pluess et al. 2012a, b).
Our aim was to make the Lists on the one hand relatively comprehensive but on the other hand simple enough for later implementation into policy tools. Such an approach was reflected in the composition of the Watch List. It contains species that are not present in the Czech Republic but require attention (because they are already established and cause impact in the neighbouring countries or areas in Europe with similar climatic conditions, and their import is highly probable), but also species already present in the Czech Republic, but currently still restricted to cultivation, captivity or another kind of controlled environment. This allows for raising attention to those "knocking on the door" as well as those already cultivated/farmed species which should be monitored.
Implementing the Black Lists into legislative tools in the Czech Republic is, as in many other countries, constrained by limited integration of IAS-related agendas among different sectors and individual concerned bodies (e.g. nature protection, agriculture, forestry, aquaculture and fishery, hunting, pet industry and trade with various species and products, research, municipalities etc.). In the Czech Republic, the issue of IAS falls within the competence of the Ministry of Environment, but some activities which can on the one hand promote IAS (e.g. biofuel plants, horticulture), or on the other hand control them (e.g. phytosanitary and veterinary measures) are under the competence of other sectors, primarily the Ministry of Agriculture. Unfortunately, due to the different interests of each sector, cooperation between them is not very effective at present. These different interests lead to the inconsistency and weakening of the legislative instruments, unclear competences in the field of IAS, as well as to their ineffective management. Therefore, an essential condition of any progress in the Czech Republic is to communicate the goals and problems caused by IAS to the general public, stakeholders and policy makers to be able to successfully incorporate the legislative measures, and preventive and control management. Implementation of the new EU Regulation will significantly facilitate this process.
The lists presented here are the first attempt to provide basis for setting the priorities of policy and nature protection at the national level in the Czech Republic. The lists should also serve as a national starting point for discussion on priority IAS species at the EU level, based on the new EU Regulation on IAS (Caffrey et al. 2014, European Commission 2014. As the EU List has to take into account interests of individual member states, it will likely reflect to a large extent political interests rather than purely scientific assessment. Therefore national lists may provide a more flexible and effective way of dealing with invasive species. Compared to other existing Black and Grey Lists for other European countries , Gederaas et al. 2012, Nehring et al. 2013, our approach also takes into account invaded habitats and feasibility and meaningfulness of potential management; we believe that such a methodological approach to prioritization of species represents important advancement, transferable to other regions in Europe and elsewhere.

S. lanceolatum)
Released/sponta- yield (Oerke 2006) or by endangering human health (Reinhardt 2003, Salo et al. 2011. Control and eradication of IAS is of increasing importance for diversity conservation and environmental health (Pyšek et al. 2007, Smith et al. 2013. Common ragweed (Ambrosia artemisiifolia) is an annual IAS, growing on disturbed sites like roadsides, fields, riversides and gardens. It is feared for the allergenic properties of its pollen, as well as a weed in agriculture, in both instances related to high financial costs (Coble et al. 1981, Buttenschøn et al. 2009, Rosenbaum et al. 2011, Smith et al. 2013. A. artemisiifolia is native to North-America and currently spreading through Europe and Asia (Kazinczi et al. 2008). In Europe, preferred habitats are summer crop fields in summer warm climates, but also ruderal places and roadsides.
The plant reproduces exclusively by seeds. One individual can produce up to 62000 seeds in North-America (Dickerson and Sweet 1971) or up to 18000 in Europe (Fumanal 2007). Ragweed seeds can enter primary dormancy and germinate next spring, or enter secondary dormancy after failure to germinate in spring (Bazzaz 1970, Baskin andBaskin 1980) and remain dormant in the soil seed bank for up to 39 years (Toole and Brown 1946). Ragweed dormancy is broken by stratification (Bazzaz 1970).
The persistent soil seed bank of A. artemisiifolia compromises the efficacy of any kind of control measure. Even if a control option succeeds in killing green plants aboveground, some part of the population remains dormant in the soil awaiting more favorable conditions to germinate. Another disadvantage of a persistent soil seed bank is that it acts as a source of further spreading of the weed in soil containments (Nawrath and Alberternst 2013, Karrer 2014). Soil is relocated from many habitats where the plant is growing, such as construction sites or roadsides to other sites. Therefore, aim of any sustainable long-term control of common ragweed should be a reduction of the soil seed bank in established populations. Milakovic et al. (2014a and2014b) and Bohren et al. (2008) found that seed production per plant could be influenced by carefully timed mowing. This study's goal is to test the effect of different cutting regimes applied for three years (Milakovic et al. 2014a) on the quantity and quality of the ragweed soil seed bank.
Regrowth of ragweed after mowing is well-documented (Barbour and Meade 1981, Bohren et al. 2005, Bohren et al. 2008, Meiss et al. 2008, Karrer et al. 2011, Patracchini et al. 2011, Simard and Benoit 2011, Tokarska-Guzik et al. 2011) and varies with season (Milakovic et al. 2014b). Timing and frequency of cutting has specific influences on the seed production of ragweed (Simard andBenoit 2011, Milakovic et al. 2014a). Higher ranked resprouts after cuts tend to produce only female flowers (Karrer et al. 2011) and, in consequence, preferably seeds that are incorporated into the soil seed bank.
Soil seed bank of plants varies by year and season. On undisturbed soil, the annual seed production of ragweed germinates to high percentages in early next spring (Dickerson 1968, Bassett and Crompton 1975, Fumanal et al. 2008, Kazinczi et al. 2008. Soil tillage incorporates new seeds into deeper layers of the soil (Buhler et al. 1997) and promotes long time persistency of ragweed seeds (Toole and Brown 1946).
The effects of different tillage systems were analyzed with respect to the composition of the soil seed bank of arable fields (Clements et al. 1996, Buhler et al. 1997, Cardina et al. 2000, Clay et al. 2006). Up to now, no study has considered the soil seed bank of ragweed for measuring the success of control options, even though the seeds in the soil make up a great portion of the population in annual weeds with a persistent soil seed bank. In this study, we used the soil seed bank of ragweed populations as long-term efficacy measure of non-chemical control options. We varied the mowing regime of ragweed roadside populations in Austria with respect to timing and frequency (Milakovic et al. 2014a) and analyzed the soil seed bank of ragweed at the beginning and at the end of the experiment.

Methods
We sampled the soil seed bank of six roadside populations in Eastern Austria before and after three years of application of management practices. Austrian arterial road verges are cut at least two times a year; a first cut in spring and a second cut between July and October. This resulted in a significant spread of common ragweed along arterial roads since 2000 (Karrer et al. 2011, Essl et al. 2009).
The cutting experiment was set up in 2009 in the heavily infested parts of Austria (Lower Austria, Styria and Burgenland) ( Table 1). All populations have been naturalized for about one or two decades before the experiment.

Experimental design:
On each site, five experimental plots were installed on continuous spontaneous populations of A. artemisiifolia with coverages ranging from 5 to 25%. The plots were arranged along a line of 100 m, adjacent and parallel to the asphaltic surface of highways or arterial roads. Each plot sized 20 × 0.5 m and received one of the following treatments (mowing regimes), as defined in Milakovic et al. (2014a): Treatment 1: not mown (control), Treatment 2: first cut before the start of flowering (the last week of June), and second cut at the beginning of seed set (second week of September). Treatment 2 resembles the common roadside cutting regime in eastern Austria. Treatment 3: first cut after the beginning of female mass flowering (third week of August), and second cut at the beginning of seed set (second week of September), Treatment 4: first cut before the start of flowering (last week of June), second cut before the onset of male mass flowering (last week of July), and third cut at the beginning of seed set (second week of September), Treatment 5: first cut before the start of flowering (last week of June), second cut after the beginning of female mass flowering (third week of August) and third cut at the beginning of seed set (second week of September).

Soil seed bank sampling
All sites have been sampled for soil seed bank before the start of the mowing experiment in spring 2009 and after three years of the experiment in spring 2012. The sampling was always performed just before or at the very start of the germination period in the field. First sampling was done in March 2009 preceding the different treatment of the plots: 20 soil cores (depth 7cm, 285cm³, equally distributed over 100m of the experiment plot) were taken at each site. After three years of applying the various treatments, in March 2012, 19 soil cores were taken per plot on each site. The soil cores were analyzed for ragweed seed content using a wet sieving machine (Retsch). We counted all intact seeds and put them into wetted Petri dishes. In order to detect the proportion of viable seeds, first germination was induced by putting them into climate chambers at the following conditions: daylight for 8 hours at 30 °C and darkness for 16 hours at 15 °C. We stopped the germination trial after 4 weeks, left the dishes for drying out and stored them for 4 weeks at +4 °C in darkness, in order to overcome secondary dormancy by additional stratification. Afterwards, a second germination period was started at the same conditions like in the first session.
All seeds that did not germinate within the second germination session were tested for vitality by a standard staining (TTC-test with 1 % solution of 2,3,5 triphenyl tetrazolium chloride in pure water). For that, Ambrosia-achenes were first imbibed in tap water at room temperature for 24 hours. The achenes were then cut open with a scalpel to expose the embryo. The bigger part of the achene was used for testing, the other part was discarded. Achene halves were put into petri dishes, covered with TTC solution and left at 30 °C for 6 hours in absolute darkness. Finally seeds were evaluated under a dissecting microscope. All fully stained seeds were classified vital.
The soil seed bank samples in 2009 were taken from the whole sites that where covered consistently with A. artemisiifolia, and can therefore be used as baseline data for comparison to the soil seed bank counting at the differently treated plots three years later. That way, it is possible to observe the effect of the tested mowing regimes on the soil seed bank after three years of application. table 1. Location (coordinate system WGS84) and habitat characteristics (road type, road orientation, initial ragweed coverage (%)) of the experimental sites along arterial roads in Austria. Data were analyzed by GLM (generalized linear model) using Poisson distribution procedures and a log link in the package Statistica 10 (StatSoft 2011). Treatment was included in the model as independent categorical factor and seed number per m 2 as dependent variable. Pairwise differences between treatments were judged at 95% confidence intervals. We compared the overall most effective treatment with the initial seed bank of the populations of each site by Kruskal-Wallis Tests.

Soil seed bank at different sites
In 2009, soil seed bank varied from 123 to 823 (522 in average) seeds per m² at all sites (Table 2), with germination rates varying from 53 to 100% (mean 80%). In 2012, soil seed bank at different sites varied from 0 to 1061 seeds per m². The germination rates were generally very high (mean 91%). From the 2012 samples, no seeds germinated during the second germination test, and no living seeds could be detected by the subsequent TTC test.

Soil seed bank in different treatments
After 3 years of applying different mowing regimes, significant differences in the soil seed bank under different treatments were found (Wald χ 2 (5) = 188795; p ≤ 0,01). The soil seed bank of treatment 1 (control, unmown) was three times higher than the soil seed bank of the population before the experiment (Figure 1). The soil seed bank of treatment 2 did not differ significantly from the soil seed bank of the population in 2009 (Figure 1). The soil seed bank of the treatments 3, 4 and 5 decreased by ca. 80%, 60% and 45%, respectively, compared to the magnitude order before the experiment (Figure 1). Efficacy of treatment 3 is obviously highest in controlling the ragweed populations sustainably. The soil seed bank decreased on all sites on the plots of treatment 3 (Figure 2), at most sites significantly (Table 3).

Discussion
The number of ragweed seeds per m 2 found in populations along Austrian roadsides before the start of treatments in 2009 indicate that those are all well-established populations that cannot be controlled by a one time management action. The aboveground assimilating part of the A. artemisiifolia population varied between the sites at the beginning of the experiment (Table 1) but showed similar dynamics to the soil seed bank towards the end of the experiment. Compared to the soil seed bank of other ruderal habitats (waste lands and set-asides) our roadside populations showed relative low seed densities. Fumanal et al. (2008) describe seed densities ranging from 510-3324 seeds per m² in the upper 5 cm of soil. This indicates that the Austrian roadside populations are relatively young but 'active' populations. Corresponding to the high population turnover rates, most seeds accumulate in the uppermost soil layer and germinate at high rates to produce many new seeds every generation. The fraction of old seeds from former population establishment phases that might have lower germination rates, seems to be relatively low as the overall germination rates of the seeds in the soil is considerably high ( Table 2). The seed bank densities of ragweed along Austrian highways are generally lower than in European arable fields (Vitalos and Karrer 2008). Habitat types that have been infested by ragweed for decades, like abandoned fields in N-America, have a load of 0-200 ragweed seeds per m² even when sampling only the persistent soil seed bank in summer (Rothrock et al. 1993). Bigwood and Inouye (1988) found on average 36 ragweed seeds per m 2 in the upper soil (0-8 cm) and 57.6 seeds per m 2 at a depth of 8-16 cm in an old field in Maryland (US). Raynal and Bazzaz (1973) counted means of 64 ragweed seeds per m 2 in maize fields on former forest soil and 4.8 seeds per m 2 on former prairie soil, when analyzing the upper soil (0-5 cm) in early spring; autumn samples did not contain ragweed seeds. Considering that the Austrian ragweed seed populations along highways are concentrated at the upper horizons of the road shoulder soil, they can be classified as very active and contribute to an increasing infestation.
Because most management options act on the green parts of the plant, they are not sustainable. The most desired aspect of ragweed control is the successful elimination of persistent seeds from the soil. The results of this long term experiment show, that the soil seed bank can be diminished vigorously by a sophisticated mowing management. The mowing regime should consist of a first cut in August, just at the first appearance of female flowers, and a second cut in early September, just before fertility of the female flowers on the regrowth from the base (Milakovic et al. 2014a). According to our results, we suggest to rate this mowing regime as the most sustainable and environmentally friendly control option, because it progressively leads to indirect depletion of the soil seed bank. This way the ragweed populations decline and can be managed easier.
Hence the biologically most effective control measure of pulling out the remaining few plants by hand (Bohren et al. 2008) might become economically feasible. We advise analyzing the soil seed bank of ragweed before installing a field experiment or defining a management regime for ragweed control, as well as after the activity. Thus sustainability can be proven. The knowledge about the status of soil seed bank is particularly important for ragweed populations growing on roadsides, as the upper soil is prone to transportation elsewhere, which contributes to further dispersal of ragweed seeds and creates new populations.

small but tough: What can ecophysiology of croaking gourami Trichopsis vittata (Cuvier, 1831) tell us about invasiveness of non-native fishes in Florida?
introduction Destructive (sometimes catastrophic) ecological impacts have been attributed to the introduction and establishment of non-native fishes across the globe (Canonico et al. 2005;Pelicice and Agostinho 2009;Vitule et al. 2009). However, the severity of negative consequences of non-native fish invasions varies greatly amongst taxa. Variation in the ability of species to establish and spread (i.e., 'invasiveness' sensu  has provided clues to underlying ecological attributes correlated with invasiveness (García-Berthou 2007). Understanding the characteristics associated with invasiveness is especially important in predicting potential establishment and spread of newly introduced species or those considered a threat for introduction. Most studies aimed at discriminating ecological features of invasive species quantify, collate and report life-history, ecophysiological, and other data for species that have become invasive (Kolar and Lodge 2002;García-Berthou 2007). Less abundant are data on species that were introduced and died out over time, or those that were introduced and established but did not become invasive. Data on those non-invasive species can be difficult to obtain when species were not intentionally introduced (e.g., via stocking). Population dynamics of fishes that were not introduced intentionally (e.g., aquaculture escapes) may not be closely monitored. Nonetheless, the fate of these populations is important as they may provide clues to allow researchers to be able to identify characteristics unique to invasive fishes from those shared between invasive and non-invasive species.
In Florida, there are dozens of non-native fish species that have established and spread widely within the state, especially in the southern half of the peninsula. However, a few species have established but remain localised (Shafland et al. 2008;USGS-NAS 2014). Croaking gourami Trichopsis vittata (Cuvier, 1831) was first collected in 1978 and was considered extirpated in the 1990s; however, a localised population was rediscovered in 2013 (Schofield and Pecora 2013; Fig. 1). The species may have persisted in a relatively small area for several decades where it escaped detection. Failure of a species to spread widely after establishment may be due to many factors, such as ecophysiological intolerance or biotic interactions with predators and/or competitors. Little is known regarding the ecophysiology of T. vittata, other than the fact that it is an air-breather, making it capable of living in anoxic waters. Other ecophysiological attributes (e.g., tolerance to salinity, extreme temperature) were unknown before this report. Herein, we investigate two ecophysiological parameters for T. vittata thought to be conducive to invasiveness in Florida (cold-and salinity-tolerance). We compare those (along with hypoxia-tolerance) to published reports for other non-native fishes with much larger geographic ranges within the State. We ask: Can T. vittata's small geographic range be explained by its relative lack of ecophysiological 'toughness' (i.e., ability to withstand environmental extremes)? In other words, are fishes more tolerant to cold temperatures, low oxygen and salinity predicted to have larger invasive geographic ranges? We hypothesised that T. vittata's small non-native range could be related to a lack of tolerance of ecophysiological variables, and expected it to be less tolerant to environmental variables than sympatric non-native fishes with large ranges.

Methods
Specimens of T. vittata were collected with dip nets in March and April 2014, from Loxahatchee National Wildlife Refuge, Florida, USA. Fish were transported to the USGS laboratory in Gainesville, Florida within 48 hours of capture. Upon arrival, fish were treated with Pond Rid-Ich® Plus™ (Kordon LLC, Hayward, CA, USA) and erythromycin antibiotic. In the laboratory, fish were held indoors in 380 L fiberglass tanks with aerated well water (0.2 psu, hereafter termed "0" psu, 21-25 °C) and were fed daily with commercial flake food. Individuals were held in these conditions for about one month before experiments began. Before each experiment, fish were measured (± 0.1 cm total length [TL]), weighed (± 0.1 g) and placed into individual plastic bins (17 × 14 × 11 cm) filled with 8 cm of well water. Bins were equipped with small plastic plants, lids to prevent escape and were blinded on three sides to prevent fish from seeing each other. Because gourami are obligate air breathers, no air was provided except for the low-temperature tolerance experiment, where an airstone was placed in each bin to aid in mixing of the water for even temperature distribution throughout the bin. For both salinity experiments, fish were kept in individual bins inside a temperature-controlled room set at 26 C for the duration of the experiments. Salt water was pre-mixed to various salinities using well water and aquarium salt (Crystal Sea® Marinemix, Marine Enterprises International, Baltimore, MD, USA) before water changes using YSI Professional Plus Multiparameter meter (YSI Inc., Yellow Springs, Ohio, USA; ± 0.05 psu). For all experiments, when death was confirmed, water temperature was measured using a hand-held digital thermometer (EXTECH® waterproof thermometer model #39240, EXTECH Instruments®, Nashua, NH, USA). Individual fish were used only once in one experiment (low-temperature tolerance, chronic salinity-tolerance or acute salinity-tolerance).

Low temperature tolerance
The low-temperature tolerance experiment was conducted in April 2014 inside an environmental chamber in which temperature could be controlled by continuously decreasing the air temperature at a constant (i.e., linear) rate. Two endpoints were determined: loss of equilibrium and death (i.e., lower lethal limit). Loss of equilibrium (LOE) was defined as the fish's inability to right itself after being gently prodded, and death was defined as the extended lack of movement by the fish after it was gently prodded while in the water. Twenty-five fish were used and were not fed during the duration of the experiment. Fish were placed in the environmental chamber in individual bins and left undisturbed for 72 hours at 24 °C to acclimate. The experiment began by decreasing the air temperature by 1 °C hr -1 to produce an equivalent decline in water temperature. Control fish (n = 5) were immediately moved from the environmental chamber to a stable "warm room" at 24 °C (± 1 °C) until the end of the testing period, when all experimental fish (n = 20) had succumbed to death. Each hour, air temperature in the chamber was manually adjusted to produce a constant decrease of water temperature at the rate of 1 °C per hour. Temperature of each bin was measured with a hand-held digital thermometer every 20 minutes. All fish were checked for LOE and death every 20 minutes; time and temperature were recorded when LOE and death were confirmed.

Chronic salinity tolerance
An initial pilot study was conducted on n = 20 individuals to establish a general range of salinity tolerance and determine what experimental salinity levels would be used for the experiment. For the chronic salinity tolerance experiment, fish were allocated randomly to one of five treatments (0 [control] = 8 replicates, 20 psu = 11 replicates, 22.5 psu = 12 replicates, 25 psu = 12 replicates, 27.5 psu = 13 replicates). Fish were held initially for 48 hours in well water after which salinities were gradually increased at a rate of 2.5 psu every 2-3 days (5 psu per week) until fish reached the predetermined target salinity. Once the last experimental fish reached its target salinity, all fish remained in their respective salinities for an additional 30 days or until death. Fish reached their target salinities in a staggered (time-wise) fashion; however, each time the salinities were changed in one or more of the treatments, water changes were performed for all of the fish (including controls) to maintain similarity of handling across treatments. Fish were fed twice per week with a mixture of flake food and pellets on days before water changes. Fish were checked 1-2 times per day, seven days per week for death.

Acute salinity tolerance
To determine how T. vittata responded to acute salinity changes, fish were transferred directly from well water (0 psu) into various salinity treatments: 0 [control] = 8 replicates, 14, 16, 18, 20 psu = 10 replicates each. Similar to the chronic-salinity tolerance experiment, values for salinity treatments were derived from a pilot study. After being transferred to their respective treatments, fish were left in bins for seven days or until death. Fish were checked for mortalities every hour for the first six hours and then once per day for the remainder of the experiment.

Analyses
Cold-tolerance of T. vittata was compared to published values for other previously tested non-native fishes. We statistically compared four species that are sympatric with T. vittata (e.g., are found in Loxahatchee NWR) and are widely distributed across south Florida ( Fig. 1). These species were tested in our laboratory using the same technique, acclima-tion temperature, experimental equipment and rate of temperature decrease used here for T. vittata Schofield and Huge 2011;Schofield unpub. data). We only compared data for individuals tested in freshwater and acclimated to 24 °C. Mean temperature at death (lower lethal limit) for these species was compared with one-way analysis of variance (ANOVA), and Dunnett's T3 post-hoc test was used to discriminate homogeneous subsets. Levene's test was used to test for heteroscedasticity.
One-way ANOVA was used to compare fish mass among salinity treatments, and Levene's test was used to check for heteroscedasticity. Life expectancy was estimated with the Kaplan-Meier product-limit estimator (Kaplan and Meier 1958) and the log-rank test was used to compare survivorship curves (Savage 1956;Cox and Oakes 1984). For the acute-salinity challenge, all treatments began at the same time (time = 0). However, for the chronic-salinity experiment, fish reached their target salinities sequentially (i.e., staggered over time). Thus, for the chronic-salinity experiment the day the fish reached their target salinity was designated as time = 0 for that treatment. We set our alpha level for statistical significance at 0.05. All data were analysed using SPSS version 13.0.
For the chronic salinity-tolerance experiment, fish mass averaged 0.93 g (+ 0.28 SD; range 0.30-1.5 g; n = 56) and mean TL was 4.2 cm (+ 0.55 SD; range 3.0-5.1 cm; n = 56). Fish mass did not vary significantly by treatment (one-way ANOVA F = 0.11, df = 4, P = 0.58). At the end of the experiment, survival was 100% at the control salinity (0 psu), 63% at 20 psu, 25% at 22.5 psu, and 8% at 25 psu (Fig. 3a). All fish at 27.5 psu died by the 24 th day after reaching their 27.5 psu salinity goal. Because the majority of the data for the 20 psu treatment was censored (i.e., the majority of fish in this treatment survived the challenge), it was not possible to compute a survival estimate. Mean (95% Confidence Interval [CI]) survival estimates for other treatments are: 18 days (11-25) at 22.5 psu, 10 days (4-16) at 25 psu, 7 days (3-11) at 27.5 psu. All treatments were significantly different from the control except 20 psu (although the P-value was close to significance; log-rank statistic = 3.41; P = 0.065) The mean mass of fish used in the acute salinity-tolerance experiment was 0.81 g (+ 0.31 SD; range 0.30-1.6 g; n = 48) and mean length was 4.1 cm TL (+ 0.54 SD; range 3.0-5.2 cm). Fish mass did not vary significantly by treatment (one-way ANOVA F =  2. Mean temperature (+ 2 SE) at which fishes died in cold-temperature tolerance experiments (i.e., lower lethal limit). Letters denote significant differences (one-way ANOVA with Dunnett's T3 post-hoc test; see text for details). References for data sources are given in Table 2. 0.98, df = 4, P = 0.43). After the acute salinity change, T. vittata at 20 psu exhibited 60% mortality within the first four hours and 100% mortality within the first six hours (mean survival estimate = 4.5 hrs; 4.1-4.9 hrs 95% CI). The 18 psu treatment group displayed 70% mortality after 24 hours, with no fish surviving longer than 48 hours (mean survival estimate = 30 hrs; 21.5-38.5 95% CI; Fig. 3b). At salinities of 0 and 14, survival was 100% and at 16 psu, it was 90% at the end of the experiment. No survival estimates were calculated for these three treatments as survival was so high (and subsequently most of the data were censored). Survival was equivalent for 0 and 14 psu (100%) and did not differ significantly between 0 and 16 psu (log-rank statistic = 0.80; P = 0.37) nor 14 and 16 (log-rank statistic = 1.00; P = 0.32).

Discussion
Trichopsis vittata has been known from Florida since the 1970s, when an established population was discovered within 10 km of its current range (Courtenay et al. 1984(Courtenay et al. , 1986Schofield and Pecora 2013). Its introduction source is unknown; however, at the time of discovery it was speculated that it had escaped from nearby aquarium fish farms (Courtenay et al. 1984(Courtenay et al. , 1986. Over time, the species was thought to have been extirpated (Shafland 1996;Shafland et al. 2008) until its recent rediscovery (Schofield and Pecora 2013). No fish-monitoring programmes cover urban areas in this region of Florida, so it is unclear how long the fish was established before its recent collection at Loxahatchee NWR. Furthermore, it is unclear whether the species had died out and was subsequently re-introduced or whether this is a remnant population. Nevertheless, its ability to persist in this small range for many decades makes it an interesting candidate for study. The purpose of this investigation was to document ecophysiological attributes of the species and qualitatively compare them to sympatric species, to see if perhaps reduced ecophysiological tolerance might be related to the lack of geographic expansion. In general, our hypothesis (low ecophysiological toughness ≈ small geographic range) was not supported. Ecophysiological traits of T. vittata and nine sympatric nonnative fishes known from Florida freshwaters were tabulated (  Fig. 1). Tolerance of these species to hypoxia and low-temperatures was graphically compared (Fig. 4). Trichopsis vittata was more tolerant of cold than many sympatric non-native fishes, leading us to believe it could tolerate habitats north of its current range; however, it has not expanded its range in any direction. Furthermore, its ability to breathe atmospheric air (via a labyrinth organ) imparts an ability to live indefinitely in water devoid of oxygen. It should be able to tolerate a variety of marginal habitats such as shallow pools, vegetation-choked swamps, and habitats with low light levels as it does at Loxahatchee NWR and in its native range (Rainboth 1996). As for salinity, we documented herein that T. vittata was tolerant to acute shifts in salinity to 16 psu and gradual shifts to 20 psu. This level of tolerance is lower than published values for most cichlids, but greater than many non-cichlid invasive fishes (Table 2). Nonetheless, it is a species that is probably tolerant enough to occupy freshwater tidal or low-salinity estuarine areas, or use them as salt bridges for dispersal. Yet it has not moved into coastal areas even though the current population is less than 20 km from the Atlantic coast. In summary, while T. vittata seems as tough or tougher than other sympatric non-native fishes (in terms of ecophysiology; Table 2, Fig. 4), it has not been able to capitalise on these advantages and expand its geographic range as the others have. It is possible that tolerances to environmental parameters are not directly correlated with geographic range for this group of species, but instead serve as air-breather unknown 10 psu (Capps et al. 2011) 9-11 for P. multiradiatus; 4-6 for P.
There are many other factors that could explain the lack of geographic range expansion for T. vittata. Some of the most obvious candidates include body size, diet and their interaction. Trichopsis vittata is smaller than other sympatric non-native fishes and occupies a relatively low position on the predation spectrum (i.e., primarily consumes small invertebrates ;Rainboth 1996). This combination of attributes separates T. vittata from the other non-native fishes that are either: 1) large-bodied species that consume benthic algae and detritus (e.g., Pterygoplichthys spp., O. aureus) or 2) largeto medium-sized fish predators (cichlids, C. batrachus). One species that does not fit this pattern is H. letourneuxi, which consumes both invertebrates and fish and does not reach a large body size (Table 2), yet is extremely invasive (Kline et al. 2013; Fig. 1). Protection from bony dermal plates may confer an additional advantage to catfishes (H. littorale, Pterygoplichthys spp.) and bolster their ability to spread geographically. Other factors that could affect invasiveness include biotic resistance (e.g., Thompson et al. 2012), specific requirements for egg/larval development or nesting, multiple introductions (Collins et al. 2002), predation susceptibility (e.g., Rehage et al. 2009) and propagule pressure (Colautti et al. 2006). Application of modern modelling techniques may allow researchers to identify which variables are most important for an invader's success (and spread) and the critical thresholds for those variables (e.g. Kolar and Lodge 2002).

Distribution and abundance of exotic earthworms introduction
Pleistocene glaciations extirpated native earthworms from much of North America, leaving landscapes devoid of earthworms until the introduction of exotic earthworms (Oligochaeta; Lumbricidae) during European settlement , 2006, Holdsworth et al. 2007a, 2007b. The effects of exotic earthworms on forest ecosystems are well documented , 2006, Holdsworth et al. 2007a, 2007b and vary by feeding strategy. Leaf litter-dwelling, small-sized epigeic species are least destructive, consuming and mixing the top organic layers into textured, homogeneous litter. Endogeic species burrow through the top soil horizon; their physical effects on ecosystem ecology are greater than epigeic worms but less than anecic worms. Anecic earthworms penetrate deep into the soil, transporting surface litter into the mineral layer (Addison 2008) and increasing soil porosity and water infiltration (Anderson 1988). Removal of leaf litter and deposition of casts on the soil surface by anecic earthworms can also increase soil erosion and nutrient run-off (Edwards and Bohlen 1996).
Material transport by anecic worms, their large adult size, and dense populations have led to substantial ecosystem changes in some parts of North America . Earthworms can accelerate litter decomposition , Suárez et al. 2006, Holdsworth et al. 2007a, 2007b, Addison 2008) and reduce plant species richness , Holdsworth et al. 2007a, 2007b. Suárez et al. (2006) found that litter remaining in earthworm-invaded areas in New York was 30-60% less than in reference plots. Holdsworth et al. (2007a) found in a Wisconsin forest that exotic earthworms reduced plant species richness in heavily invaded plots by 17%. Similarly, Hale et al. (2006) documented a negative relationship between exotic earthworm diversity and plant diversity in a Minnesota hardwood forest.
Most studies of exotic earthworms have occurred in temperate regions (Hale et al. 2006, Suárez et al. 2006, Holdsworth et al. 2007a, 2007b, Addison 2008; less is known about the distribution and effects of earthworms in subarctic boreal forests (Cameron et al. 2007, Cameron and Bayne 2009, Sanderson et al. 2012. In northern Alberta, Cameron et al. (2007) found boat launches and roads had the highest probability of earthworm occurrence. Their results suggested vehicle transport and bait abandonment as primary mechanisms of earthworm introduction.
As for most invasive species, human activities, particularly road construction and unintentional transport, likely increase the rate of spread for exotic earthworms above their natural dispersal rate of 5-10 meters a year , Addison 2008). Consequently, exotic earthworms more likely occur near roads due to availability of habitats disturbed by road construction and maintenance that allow for potential establishment, as well as the creation of dispersal corridors (Cameron et al. 2009). Vehicles themselves function as dispersal vectors for earthworm cocoons, which are sticky, mucus coated sacks containing developing embryos . Several species such as Lumbricus terrestris (anecic) and Lumbricus rubellus (epi-endogeic) are sold commercially as fishing bait and are possibly introduced into ecosystems when anglers discard unused bait (Cameron et al. 2007).
Seventeen species of earthworms are known to occur in Alaska (see records in Gates 1972, 1974, Reynolds et al. 1974, Reynolds and Wetzel 2008, Reynolds 1977, 1980, Berman and Marusik 1994, Costello et al. 2011, Rinella et al. 2014 material 1: Alaska earthworm records). Of these, 14 are exotic worms introduced from the Palearctic and have become established. Eisenia fetida (Savigny, 1826), a Palearctic species, is commonly used for indoor vermicomposting in Alaska, but due to its low cold tolerance (Greiner et al. 2011, Meshcheryakova andBerman 2014), it is unlikely to become established in Alaska. Two species of earthworms found in southeast Alaska (Arctiostrotus sp. and Sparganophilus sp.) may be native to Alaska or may have been transported from elsewhere in North America.
Factors such as pH and temperature likely limit earthworm distribution, especially in boreal regions like Alaska (Chan andMead 2003, Addison 2008). Earthworms are usually associated with soil pH of 5-7.4, although D. octaedra inhabits soil pH as low as 2.8-3.6, and L. rubellus has been found in areas with pH ≥ 3.0 (Addison 2008). Survival of earthworms in low temperature areas depends on the species and stage of development (Greiner et al. 2011, Meshcheryakova et al. 2014). Meshcheryakova and Berman (2014), by comparing cold hardiness and present distributions of earthworm species in Siberia, concluded that varying cold tolerance of the species considered contributed toward their present distribution ranges.
A rapidly warming climate in Alaska is likely improving environmental conditions for earthworms. Wetlands in Alaska are warming and drying (Klein et al. 2005, Riordan et al. 2006, Berg et al. 2009) and average winter temperatures have warmed 3.5 °C in the last 50 years (Karl et al. 2009). Drying wetlands and warmer winters may provide increasingly suitable habitat for exotic earthworms. Addison (2008) suggested that even small increases in winter temperatures will lead to large increases in earthworm habitat.
The present study documents species composition, distribution, and habitat correlates for earthworms in the Kenai National Wildlife Refuge, a conservation area in southcentral Alaska. A secondary goal is to examine relationships between earthworm occurrence and distance from human-disturbed areas, such as roads and popular fishing areas. The final goal is to identify factors potentially limiting earthworm distribution, such as pH and soil moisture, which are likely to change as the climate continues to warm on the Kenai Peninsula.

Study area
Located on the Kenai Peninsula in southcentral Alaska, USA (60°N, 150°W), the Kenai National Wildlife Refuge (KNWR) covers 777,000 ha. Mountains and glaciers characterize the southeastern KNWR (Figure 1). The Kenai Lowlands, mantled by glacial deposits that vary in texture and are capped by silt loam derived from post-glacial windblown loess, cover the western portion of KNWR. The Lowlands consist of wetlands and mixed boreal forest (Klein et al. 2005) dominated by black spruce (Picea mariana), white spruce (Picea glauca), birch (Betula papyrifera), and quaking aspen (Populus tremuloides). The climate is boreal with a maritime influence. Temperatures are rarely greater than 26 °C in summer or less than -18 °C in winter. The frost-free growing season varies from 71-129 days depending on location, with about 480 mm of total precipitation per year (U.S. Fish & Wildlife Service 2010).
While most of the KNWR is currently managed as congressionally designated Wilderness, over 130 historic cabins have been inventoried on the Refuge along with other historic resources associated with mining, trapping, oil development, and road construction, mostly in the north of the refuge. Commercial mining and fishing in the area occurred in the late 1800s and early 1900s. Oil exploration began in the northern part of KNWR in the Swanson River area in the mid-1950s, resulting in 2,900 km of bulldozed seismic lines. There are also 240 km of utility and transmission lines and 180 km of established trails within the Refuge (U.S. Fish & Wildlife Service 2010). Fire and associated suppression activities have also been prevalent within the refuge in the past. Major fires of unknown origin occurred in 1871, 1883, 1891, and 1910. Two large, human-caused fires (1947: 125,000 ha and 1969,000 ha) resulted in replacement of mature spruce forests by a mosaic of young mixed conifer-deciduous forest in various stages of succession. Many historic disturbances provided opportunities for earthworm introduction and establishment on the KNWR.

Experimental design
Earthworms were sampled throughout the KNWR during July and August 2011 at 69 total sites representing three levels of human impact. These levels of human impact were characterized by explicit vectors of introduction: (1) boat launches (n = 20), (2) road corridors (n = 20); and (3) low impact areas (> 600 m from any road or facility and 50 m from any trail or river; n = 29). The sampling site locations were chosen within a GIS (ArcGIS v.10.1) by first identifying suitable areas for each impact level and then randomly selecting sample sites. Three 0.09 m 2 plots were established at each site to sample earthworm occurrence. At road and boat launch sites, plots were placed two, three, and five meters from the road or edge of a boat launch. The plots were located approximately 10 m apart when possible, establishing a wider area for detecting earthworm occurrence. This protocol was modified at four of the low impact sites accessed by float plane, where only a single 0.09 m 2 quadrat was sampled at each site.

Plot level sampling
We sampled earthworm abundance at each plot using a 30 cm × 30 cm quadrat. Within each quadrat we removed and hand-sorted surface organic material for earthworms. We extracted additional earthworms with a liquid mustard solution of 40 g ground mustard powder in 3.8 L water (Lawrence and Bowers 2002). Earthworms were collected and stored in 70% ethanol. Specimens were deposited in the entomology collection of the Kenai National Wildlife Refuge (coden: KNWR) and specimen data were made available via Arctos (http://arctos.database.museum/).
We identified earthworm specimens to species level when possible based on visual observations of external morphology (Reynolds 1977). Juveniles were grouped into two categories: (1) Lumbricus spp. and (2) other immature. In addition, six specimens collected at geographically remote sites were identified to species level using the mitochondrial COI barcoding gene to confirm taxonomic identification using the Basic Local Alignment Search Tool (BLAST) and the BOLD identification engine (http://www.boldsystems.org). We submitted sequence data to BOLD where they are publicly available.
We measured leaf litter depth by clearing a small area and measuring the vertical depth of the leaf layer with a 30 cm ruler within each plot. Soil pH, (Soil pH Meter, HANNA, RI), and moisture (Digital Moisture Meter, General, NY) were measured in the field. In each plot, we estimated the percentage cover of litter, grass, forbs, moss, and lichen. General forest type (deciduous, mixed, conifer) of the area was determined from field observations and a GIS land cover layer.
We estimated ash-free dry biomass (g) from the length (mm) of each preserved specimen with the allometric equation of Hale (Hale et al. 2004): g= e 2.2853ln(mm)-11.9047 Hale (2004) found that the allometric equations for Octolasion tyrtaeum, Lumbricus spp., and Dendrobaena octaedra were not significantly different from one another, allowing one equation for all species.

Lumbricus spp. distribution sampling
Anecic Lumbricus species are potentially more damaging than other genera (Eisenhauer et al. 2007). Consequently, at sites where Lumbricus spp. were found, we used three transects to estimate the extent of local distribution. One transect was perpendicular to the initial site and the other two at approximately 45 degree angles from the site. At 10 m intervals along each transect we sampled three plots for earthworm presence using liquid mustard extraction within a 0.25 m 2 quadrat. To delineate the boundaries of this infestation, we continued sampling until we failed to find earthworms in all three plots at two consecutive 10 m intervals. Earthworms were collected and stored in 70% ethanol and later identified in the lab.

Statistical analysis
Individual plots served as replicates for each site, but we used site level data for most analyses by averaging plot level data. A site was considered to contain earthworms if individuals were detected in at least one plot. Soil moisture and leaf litter depth were square-root transformed for normality. We calculated remoteness for each site as a measure of distance from the nearest road. This distance, Y (m), was calculated in GIS by using true surface distance, as it accounted for elevation changes and also masked out lakes. To approximate a normal distribution for analyses, we transformed this distance using ln (Y + 1).
All analyses were performed using R, version 3.1.2 (R Core Team 2014). We assessed independence of earthworm presence and human impact level (road, boat launch, low impact) and vegetation type (conifer, deciduous) using chi-square tests of independence. Correlations of independent variables were examined with the corr.test function of the psych package, version 1.5.1 (Revelle 2015), accepting default arguments.
Prior to occupancy modeling, principal components analysis (PCA) was used to reduce the dimensionality of the eight habitat variables soil pH, soil moisture, leaf litter depth, and moss, grass, lichen, litter, and forb cover percentages. PCA was performed using the PCA function from the FactoMineR package (Husson et al. 2015), version 1.29. We used the estim_ncp function, also from the FactoMineR package, to determine the optimal number of dimensions to use in the PCA.
Results of the PCA were included in occupancy models (MacKenzie et al. 2003(MacKenzie et al. , 2006 using the occu function of the unmarked package, version 0.10-4 (Fiske and Chandler 2011). Detection probability was assumed to be constant. Site occupancy was modeled using all 2 5 =32 permutations of first-order terms for impact level (factor with three levels), forest type (factor with two levels), distance to roads, and values from the first two PCA components. Finally, we obtained parameter importance and AICc model-averaged estimates of coefficients of the independent variables from the full set of candidate models using the importance and modavg functions from the AICcmodavg package, vesion 2.0-3 (Mazerolle 2015).
We used MANOVA to determine the effects of the impact level and earthworm occurrence on the three dependent variables of soil pH, soil moisture, and leaf litter depth.

Distribution and abundance
We found three exotic earthworm species, Dendrobaena octaedra, Dendrodrilus rubidus, and Lumbricus terrestris, on KNWR. We failed to detect Lumbricus rubellus, known from one location on KNWR (http://arctos.database.museum/guid/KNWR:Ento:6755), even though one of our sampling sites was only ~ 48 meters from this known occurrence. Specimen records are included in Suppl. material 2: Specimen records.
The six individuals that were genetically identified using the COI gene showed > 96% probability of identity to their respective species based on a BLAST results. BOLD process ID's for sequenced specimens are also included in Suppl. material 2: Specimen records.
No site had more than two species confirmed as present. Only four sites (three boat launch sites one road site) contained two species of earthworms, while the majority of sites contained only one species. Dendrobaena octaedra was the most widespread, occurring at 48 (70%) of 69 sites. Most immature earthworms appeared to be D. octaedra based on morphology. Dendrodrilus rubidus occurred at two sites geographically distant from one another. Lumbricus terrestris occurred at three sites, all of which were boat launches located adjacent to one another in the northern part of the Refuge (Figure 1). Along transects surveyed at these sites, we found that L. terrestris had dispersed only 30 to 110 m from the boat launches.
Overall, earthworms occurred at 49 (71%) of the 69 sampled sites. Nearly all road sites (18 of 20 total sites, 90%) had earthworms in at least one plot, as did most boat launches (17 of 20 total sites, 85%). In contrast, only half (14 of 29 total sites, 48%) of the low-impact sites contained earthworms.
Earthworms occurred more frequently at roads and boat launch sites than expected, but much less frequently at low impact sites than expected (Table 1, chi-square test of independence, χ 2 = 11.18, df = 2, p = 0.004, n = 69 observations). Earthworms were found more frequently than expected at sites dominated by deciduous trees and shrubs and less frequently than expected at conifer-dominated sites (χ 2 = 13.3, df = 1, p = 0.0003, n = 65, Table 2).
Where earthworms were present, the mean density of earthworms was (28.1 ± 4.4 individuals/m 2 ), with mean densities ranging from 23.9 ± 4.5 at road sites to 33.1 ± 6.1 at boat launches (Table 3). At the three sites where transects were surveyed for L. terrestris, the mean density of this species was 37.4 ± 7.0 individuals/m 2 . Logtransformed total earthworm densities (excluding absences) did not differ significantly among impact levels (one-way ANOVA: F = 1.57, p = 0.219). Where earthworms were found, ash-free dry biomass (AFD) of earthworms showed moderate variation (0.36 ± 0.14 AFD g/m 2 , n = 49; Table 3) and was greatest at boat launches due to the presence of Lumbricus (0.65 ± 0.35 g/m 2 , n = 16), lowest at roads (0.11 ± 0.04 g/m 2 , n = 18), and moderate at low impact sites (0.19 ± 0.06 g/m 2 , n = 14). At the three sites where transects were surveyed for L. terrestris, the mean AFD of this species was 4.2 ± 1.8 g/m 2 . Log-transformed total earthworm biomass did not differ significantly among impact levels (one-way ANOVA: F = 0.818, p = 0.448). In summary, we found the highest density and biomass of earthworms at boat launches, and the least of both abundance metrics along roads.

PCA and occupancy modeling
There were many significant correlations among the habitat variables (Table 5). Depending on the method used for determining the best number of principal components to include, the optimal number was estimated to be two or three. We chose to include two components because of the relatively small sample size of our dataset (n = 65). The first two principal components that emerged from the PCA together accounted for 63% of the total variability in the original data. The first component accounted for 35% of the variability with positive loadings from soil pH, soil moisture, and grass cover and negative loadings for moss, lichen and forb cover. The second component accounted for 28% of the variability in the original variables with positive loadings from soil moisture and grass cover and negative loadings from leaf litter cover and leaf litter depth ( Table 6).
The model-averaged overall estimates of occupancy and detection probability of earthworms were, respectively, 0.83 ± 0.07 and 0.68 ± 0.04. Among the impact levels, the occupancy estimate was highest at road sites (0.90 ± 0.09) and lowest at remote sites (0.73 ± 0.16), although confidence intervals of occupancy at all three human impact levels overlapped considerably (Table 7). Earthworms were more likely to occupy hardwood-dominated sites (0.91 ± 0.06) than conifer-dominated sites (0.47 ± 0.16).
The top-ranked occupancy model had a weight of 0.31 and included terms for forest type and distance to roads (Table 8). The second-ranked model, with a weight of 0.16, included terms for forest type and the impact levels boat launch and road distance. The evidence ratio between these two models suggested the highest-ranked model was 1.9 times more likely to be the most parsimonious model than the secondranked model, but a Δ AICc < 2 indicated that the two models were nearly equivalent (Symonds and Moussalli 2011). In fact, the combination of the road and boat launch terms, both reflecting categories of sites very close to roads, conveyed much of the same information as the distance term.
Forest type was included in all highly-ranked models (importance value of 0.98, Table 9, Table 9). Its value as parameterized (conifer as intercept, hardwood as dummy variable) was consistently positive (95% CI: 0.76, 4.14), meaning that earthworms were more likely to occur under hardwoods than under conifers. Distance to roads was the only other parameter with an importance value greater than 0.5. Even though its model-averaged 95% confidence interval included zero, the parameter estimate for distance to roads was negative in all models in which it was included, indicating that the likelihood of earthworm occurrence decreased with increasing distance from roads.

Effects of earthworm presence on soil properties
The presence of earthworms did not affect leaf litter depth or other measured soil properties. In a two-way MANOVA of the three factors impact level, forest type, and earthworm occurrence on the three dependent variables of soil pH, soil moisture, and leaf litter depth, the combined dependent variables differed among impact levels (Pilliai's trace = 0.804, F= 11.7, p < 0.001) and between forest types (Pilliai's trace = 0.243, F = 5.47, p = 0.002), but the combined variables did not differ between sites where earthworms were present or absent (Pilliai's trace = 0.037, F = 0.648, p = 0.588). Follow-up univariate ANOVA tests and Bayesian model averaging confirmed that neither soil pH, soil moisture, nor leaf litter depth were affected by the presence or biomass of earthworms.

Discussion
Exotic earthworms were found to inhabit 90% of road corridors and 85% of boat launch sites, but only 48% of low impact sites. These results suggest that human traffic influences earthworm presence in the KNWR. Similarly, Cameron and Bayne (2009) found a higher probability of earthworm occurrence at boat launches and roads compared to forest interiors and remote shorelines in Alberta, Canada. Gundale et al.  found exotic earthworms at all non-wilderness sites (fishing, timber harvest, road) in Michigan, but at only 50% of wilderness sites with no history of logging. The road system in the KNWR, while poorly developed compared to conservation units in the contiguous U.S., is fairly extensive compared to other Federal conservation units in Alaska, constituting ~1% of the refuge (100 m buffer either side of all refuge roads gives 6,420 ha). The paved 35 km Sterling Highway and graveled 31 km Skilak Lake Road together bisect the KNWR. These two unpaved roads provide connectivity to many of the 2,900 km of seismic lines (U.S. Fish & Wildlife Service 2010) and three active oil fields that have been laid down on the landscape over the past six decades. Road age has been linked to earthworm presence in northern Alberta, where Cameron and Bayne (2009) found that older road corridors (average age = 46 years) were significantly more likely to have earthworms than younger ones. The few roads on the KNWR were built in the 1950s, suggesting that they likely contributed to the dispersal of exotic earthworms, although sites on Tustumena Lake in the southern KNWR suggest that roads and survey lines are not necessary for earthworm invasion.
We found Dendrobaena octaedra to be the most widespread (adults at 70% of study sites) and abundant species (25.6 ± 4.4 ind./m 2 ) of earthworm on KNWR. This species is most likely introduced and spread by vehicles because its small size and epigeic habits (i.e., inhabit near-surface of the leaf litter) likely increase its chances of dispersal by human activities. Dendrobaena octaedra is a prominent invader throughout North America, often both the most widespread and densest exotic earthworm (Cameron et al. 2007).
Dendrodrilus rubidus was found at only two locations geographically distant from each other, suggesting independent introduction events and perhaps multiple vectors. In an unrelated sampling effort, we have also collected Dd. rubidus in the subalpine zone on the southern portion of the refuge at a site accessible only by floatplane or by foot (http://arctos.database.museum/guid/KNWR:Ento:7100), again suggesting another independent introduction.
Dendrodrilus rubidus is a fairly common earthworm that appears to be present more in northern hardwood and coniferous forests throughout Alaska and Canada (Cameron et al. 2007, Addison 2008, Costello et al. 2011) than in hardwood forests in the Midwest and other areas of the contiguous United States , Suárez et al. 2006, Holdsworth et al. 2007b). Like D. octaedra, it is tolerant of both acidic conditions and frost and, as an epigeic species, likely impacts the forest floor ecosystems less than anecic species (Addison 2008).
In contrast, L. terrestris is an anecic species that lives deep in the soil , Suárez et al. 2006, Addison 2008) and is commonly sold as fishing bait. We found L. terrestris at three boat launches within 5 km of one another, all at lakes popular for sport fishing (Figure 1). This peculiar distribution and the fact that L. terrestris is sold for bait locally (e.g, http://arctos.database.museum/guid/KNWR:Ento:6753) suggests direct bait abandonment as the main method of introduction on the KNWR. Additional species may be expected to arrive at boat launches because fishing bait can contain other species as well ).
Though we did not detect L. rubellus in our sampling effort, it is presently known to occur on KNWR at only one site, a boat launch on a popular fishing lake. As with L. terrestris, it was most likely introduced by bait abandonment.
In this study, the most important factors determining earthworm occurrence appeared to be forest type (conifer versus hardwoods) followed by distance from roads. We found that earthworms were more likely to be found at sites dominated by deciduous trees and shrubs than at sites dominated by conifers. On KNWR, conifer dominated sites tend to have acidic soils covered by a thick moss carpet, conditions unfavorable to most earthworm species. In contrast, hardwood sites tend to have less acidic soils covered by deciduous leaf litter, providing more ideal conditions for most earthworm species.
Our finding that half (48%) of the low impact sites (> 600 m from any road or facility, and 50 m from any trail or river) contained earthworms was relatively higher than Cameron and Bayne (2009), who noted 8-35% of their remote transects (300-500 m in the forest interior) contained earthworms, but similar to Gundale et al. (2005) who found 50% of wilderness areas without earthworms. The difference between occurrence patterns in the above studies and KNWR, with its remote sites far from roads supporting earthworms, is perhaps due to boat and float plane access into more remote regions of the KNWR. Similarly, Holdsworth et al. (2007b) found that of all habitat and distance variables, distance to roads was the only significant predictor of earthworm occurrence in a Wisconsin hardwood forest for most earthworm groups. Holdsworth et al. (2007b) noted that Dendrobaena species are early colonizers among earthworm assemblages. The highest occurrence of Dendrobaena near roads suggests that the KNWR may be in the early stages of earthworm colonization.
Besides surface vehicle access, other anthropogenic influences likely contribute to earthworm presence in KNWR, especially in more remote areas. These remote site invasion vectors are not easily identified. Dendrobaena octaedra was found throughout the study area, most likely introduced by road construction, but also possibly by seismic exploration, fire suppression activities, and mechanical tree crushing for moose browse in the northern part of the KNWR during the 1970s. There, extensive seismic lines, mostly in the northern part of the Refuge, have been in place since as early as the 1950s, and many remain visible on the landscape today as animal, hiking, and snowmobile trails, as well as illegal access routes for all-terrain vehicles (U.S. Fish & Wildlife Service 2010). Numerous prescribed fires and wildfires within the KNWR, together with associated control and suppression efforts using heavy equipment, provided additional opportunities for earthworm establishment.
There are also non-anthropogenic vectors that can spread earthworms such as birds (D. Saltmarsh, pers obs.) and streams. In southeast Alaska, Costello et al. (2011) found that earthworms appear to disperse along streams. They showed that several earthworm species could survive ≥ 6 days submerged in a stream.
Factors such as soil pH likely also limit earthworm distribution. Most earthworms prefer soil pH of 5-7.4 (Addison 2008). While earthworms were found in the present study at sites with slightly higher pH (5.74 ± 0.13, n = 48 sites) than sites without earthworms (5.32 ± 0.23, n = 21), the distribution observed was most likely due to the distance from human impacts rather than pH. The average pH of the low impact sites was significantly different from other site types, likely due to the high number of low Regionally in Alaska, the distribution of permafrost and cold winter temperatures, as well as soil moisture and pH, likely limit earthworms' potential Alaskan distribution. Where earthworms can survive, historic and current human activity and land use practices, and the composition of particular source populations, likely determine earthworm occurrence. The fact that all earthworm records in Alaska up to the present time have been from southern Alaska (see Suppl. material 1: Alaska earthworm records) despite rates of earthworm introductions that are likely comparable in southern and Interior Alaska indicates that the harsher, colder climate of the Interior precludes successful invasions by most earthworms. However, some of the more cold-hardy species present in the far north of the Palearctic, including D. octaedra and Dd. rubidus, may be able to survive in Interior Alaska based on the physiological and distributional data presented by Meshcheryakova et al. (2014).
Both D. octaedra and Dd. rubidus are parthenogenic, frost-hardy species, traits that, combined with their ability to tolerate acidic soils and exploit poor litter quality, contribute to their success in colonizing large areas. As with many "weedy" species, parthenogenesis facilitates rapid reproduction from very low densities, characteristic of rare dispersal events, where a single individual can establish an entire population ). The small body size of these species also facilitates spread by vectors such as tires more often than Lumbricus and other anecic species. Given its wide distribution on the KNWR and its particular ecological traits, D. octaedra will likely be able to colonize large areas of permafrost-free Alaska, an expanding region as climate warms (Osterkamp 2005). While D. octaedra has limited impacts compared to other exotic earthworm species, its presence could portend an invasion by a larger assemblage of earthworms and commensurate changes in soil properties if Alaska follows the same colonization sequence seen elsewhere in northern North America.

Conclusion
As there are no effective strategies for removing exotic earthworms once they are established, preventing invasion and slowing their dispersal are the only viable ways to reduce their overall impacts. While we found that D. octaedra was widespread, most of the KNWR was free of the more damaging Lumbricus species. Most of the KNWR can be kept free of Lumbricus species for many years due to the extremely slow natural dispersal rate of these worms. Because the main vectors of earthworms on the KNWR appeared to be vehicles and bait abandonment, logical methods for slowing the spread of species already on the landscape and preventing the introduction of additional earthworm species would be to minimize vehicular activity in areas currently devoid of earthworms and to explicitly prohibit the use of earthworms as live bait.
As the first study of earthworm diversity and distribution in the southcentral Alaska region, we established patterns of distribution likely to hold true regionally and we set the stage for considering ways to limit the further introductions of exotic earthworms in Alaska.

supplementary material 1
Alaska earthworm records Authors: Deanna Marie Saltmarsh, Matthew L. Bowser, John M. Morton, Shirley Lang, Daniel Shain, Roman Dial Data type: occurrence Explanation note: Earthworm records from Alaska exclusive of data from the present study are compiled. All literature items cited are included in the References section of the manuscript. Copyright notice: This dataset is made available under the Open Database License (http://opendatacommons.org/licenses/odbl/1.0/). The Open Database License (ODbL) is a license agreement intended to allow users to freely share, modify, and use this Dataset while maintaining this same freedom for others, provided that the original source and author(s) are credited.