A review of the non-indigenous Chinese mystery snail, Cipangopaludina chinensis (Viviparidae), in North America, with emphasis on occurrence in Canada and the potential impact on indigenous aquatic species

Evidence suggests that the Chinese mystery snail, Cipangopaludina chinensis, a freshwater, dioecious, snail of Asian origin has become invasive in North America, Belgium, and the Netherlands. Invasive species threaten indigenous biodiversity and have socioeconomic consequences where invasive. The aim of this review is to synthesize the relevant literature pertaining to C. chinensis in Canada. In doing so, we (i) describe C. chinensis ecosystem interactions in both indigenous (Asia) and non-indigenous habitats (North America and Europe), (ii) identify gaps in the literature, and (iii) determine where the species potential distribution in North America requires further exploration. We also briefly discuss potential management strategies for this species, as an aquatic invasive species (AIS), in Canada. Due to the much larger relative size of adult C. chinensis, multiple feeding mechanisms, and resistance to predation, C. chinensis can out-compete and displace indigenous freshwater gastropods and other molluscs. Furthermore, C. chinensis can affect food webs through bottom-up interactions with the bacterial and zooplankton communities by changing nitrogen and phosphorous concentrations. Also, the Chinese literature indicates the potential for C. chinensis to act as a biotransfer of contaminants between polluted ecosystems and consumers. In its indigenous range, C. chinensis was identified as a host for numerous parasites harmful to human and animal consumers alike. A comparison of the Canadian geographical distribution of reported occurrences with that for the United States indicates several potential gaps in Canadian reporting, which merits further investigation and consideration, especially in regard to federal and provincial non-indigenous monitoring and regulations. Southern Ontario had the highest number of reports that were mostly from web-based photo-supported sources. This suggests that interactive citizen science through popular apps backed by well-supported educational campaigns may be a highly effective means of tracking C. chinensis spread, which can be complementary to traditional methods using specimen-vouchered taxonomically verified natural-history collections overseen by professional curators.


Introduction
The Chinese mystery snail (CMS), Cipangopaludina chinensis, is a dioecious, freshwater, potentially invasive, viviparous mollusc introduced to continental North America around 1890 for consumption in the Asian food market (Stephen et al. 2013). The species has also been recently reported as established in the Netherlands (2007) and Belgium (2016)  ; Van den Neucker et al. 2017). Cipangopaludina chinensis is reported to be indigenous to Burma, Thailand, South Vietnam, China, Korea, Japan, the Philippines, and the Island of Java (Indiana Department of Natural Resources 2005;Li et al. 2013;Li et al. 2018; but see Fig. 1). Fecundity studies indicate C. chinensis produces shelled, fully-formed young at an estimated rate of 27-33 young per female per year (Stephen et al. 2013), although estimates of up to 100 young per female per year have been suggested (Haak 2015). The species feeds on zooplankton and phytoplankton (Indiana Department of Natural Resources 2005), either by grazing using a radula, or by filtering water (Olden et al. 2013). Cipangopaludina chinensis excrete large amounts of fecal matter, which can lead to potential re-engineering of nitrogen and phosphorous cycling within aquatic ecosystems (Olden et al. 2013). The ability of C. chinensis to resist predation, out-compete indigenous mollusc species for resources, reproduce rapidly, and shift food-webs suggests that C. chinensis is an aquatic invasive species.
In Canada, the Department of Fisheries and Oceans (DFO), Canadian Food Inspection Agency, Canada Border Services Agency, Environment and Climate Change Canada, Transport Canada, and the Department of National Defense, in collaboration with every province and territory, have a mandate to monitor and regulate invasive species in fresh and marine waters (Fisheries and Oceans Canada 2018). DFO defines aquatic invasive species (AIS) as: "Fish, animal, and plant species that have been introduced into a new aquatic ecosystem and are having harmful consequences for the natural resources in the native aquatic ecosystem and (or) the human use of the resource" (Fisheries and Oceans Canada 2018). AIS threaten indigenous biodiversity and natural ecosystems, have significant financial consequences, and can become difficult to manage once established. Globally, invasive species are second only to habitat destruction as a cause of species endangerment and extinction (Pejchar and Mooney 2009). Many AIS threaten natural biodiversity through predation, parasitism, competition, and degradation or destruction of invaded ecosystems (Fisheries and Oceans Canada 2019a). AIS may carry viruses, bacteria, and parasites and typically entail significant economic costs due to disruption of ecosystem services, impacts on aquaculture and fisheries, and damage to infrastructure (Fisheries and Oceans Canada 2019a;Pejchar and Mooney 2009). In both Canada and the United States (USA), C. chinensis is not yet federally prioritized as a species needing close management and is labelled in both countries as "non-indigenous" versus "invasive", which may lead to further confusion on the species' status (Fisheries and Oceans Canada 2019b;Kipp et al. 2020). Nonetheless, many USA states and a few Canadian provinces do recognize this species as invasive and have taken steps to prevent sale and distribution of C. chinensis. In the USA, C. chinensis is federally categorized as a non-indigenous aquatic species, although the United States Geological Survey (USGS) does offer a disclaimer that the listing information for this species is preliminary (Kipp et al. 2020).
Cipangopaludina chinensis is reported to be widespread in freshwater systems across the USA and Canada, with recent and more limited distribution in Europe (Clarke 1981;Collas et al. 2017;McAlpine et al. 2016;Jokinen 1982;Kipp et al. 2020;Stephen 2017; Van den Neucker et al. 2017). McAlpine et al. (2016) reported established populations of C. chinensis in Canada in southern Ontario, British Columbia, Quebec, New Brunswick, Newfoundland, and Nova Scotia. In the USA, C. chinensis has been reported in 32 states (Jokinen 1982;Kipp et al. 2020). The main vectors for the spread of C. chinensis in North America are human activities including boating, aquarium releases, and the addition of C. chinensis to garden ponds (Haak 2015;Havel 2011;Havel et al. 2014;Indiana Department of Natural Resources 2005;Harried et al. 2015;Solomon et al. 2010). Unlike the USA, Canada has not created a national database for C. chinensis, and there is no risk assessment or predictive species distribution model for C. chinensis in Canada. Therefore, neither the true nor predicted extent of this species distribution in Canada is known. Furthermore, the current literature dealing with the impact of C. chinensis on indigenous freshwater species, habitat preferences, and basic biological information is sometimes contradictory (Haak 2015;Olden et al. 2013;Solomon et al. 2010;Stephen et al. 2013;Sura and Mahon 2011).
Policy makers need to consider AIS effects on each invaded ecosystem in totality when developing management strategies. However, this is difficult to accomplish when information on individual AIS is spread across multiple forums (i.e., peerreviewed publications, grey literature, on-line and off-line databases, expert knowledge) and political jurisdictions. Literature reviews are therefore increasingly necessary to condense available knowledge, create conceptual frameworks of species impacts on invaded ecosystems, identify potential risk factors that could be important for future species risk assessments, and identify gaps in knowledge .
The aim of this review is to synthesize relevant literature dealing with the CMS so as to (i) describe C. chinensis ecosystem interactions in both indigenous and non-indigenous habitats, (ii) identify gaps in the literature, and (iii) determine where the species potential distribution in North America requires further exploration, with emphasis on Canada. We also briefly discuss potential management strategies for this species as an AIS in Canada.

Literature review
Internet searches for published papers written in English, French, or translated to English (often from Chinese), and in Chinese were conducted using Web of Science, Science Direct, PubMed, Google, Google Scholar, and China National Knowledge Infrastructure (CNKI). Reputable internet sources were also examined to determine the extent of information directed at efforts to enlist the public in limiting the species introduction to North American water bodies. References cited within all literature were reviewed, additional papers of interest were noted, and if relevant, included. Additionally, any works that cited these papers were reviewed for relevance. The literature searched spans the years from 1965-2019.
For C. chinensis, condensing knowledge from older literature can be problematic because the taxonomy of the CMS has been contentious, with various scientific names applied to the species over the past few decades. Regardless of the name used in each reference, we follow Integrated Taxonomic Information System (ITIS) conventions here and use "Chinese mystery snail" and "Cipangopaludina chinensis" (ITIS Taxonomic Serial No. 70329) in this review. It is noted that C. chinensis and B. chinensis are the more commonly used terms in modern English-language literature (Haak 2015;Haak et al. 2017;Johnson et al. 2009;Olden et al. 2013;Prezant et al. 2006;Smith 2000;Solomon et al. 2010;Stephen et al. 2013;Twardochleb and Olden 2016;Waltz 2008), while V. chinensis, V. chinensis malleata, Paludina chinensis, and Paludina malleata reflect an older taxonomy before C. chinensis and Cipangopaludina japonica (see further) were considered separate species (Jokinen 1982). "Vivipare chinoise" and V. malleata were used in Frenchlanguage literature for C. chinensis (Forets, Faune et Parcs Quebec 2018, Sta nczykowska et al. 1971. The names C. wingatei, C. leucostoma, and C. diminuta are synonyms in Asian literature for C. chinensis (Lu et al. 2014). Other synonyms used for C. chinensis also have included C. chinensis malleata, C. malleata, and V. chinensis (Kipp et al. 2020;Lu et al. 2014).
Taxonomic confusion between C. japonica (ITIS Taxonomic Serial No. 70332) and C. chinesis has also led to uncertainty in the North American literature with respect to both species. Cipangopaludina japonica, Japanese mystery snail, is another large, non-indigenous, viviparid (also established in North America and potentially invasive; Burks et al. 2016). Older literature has suggested this species is conspecific with C. chinensis, each representing different ecophenotypes of the same species (Jokinen 1982;Kipp et al. 2020). Adding to the confusion, C. japonica also has an array of various taxonomic names (e.g., Kipp et al. 2020;= Viviparus japonicus). However, morphologic and genetic differences between the two species have been reported, and here we follow current consensus as per the ITIS (http://www.itis.gov/) and treat both species as distinct (Chiu et al. 2002;Clench and Fuller 1965;David and Cote 2019;Hirano et al. 2015Hirano et al. , 2019Lu et al. 2014;Smith 2000;Wang et al. 2017).
While much of the older literature (pre-1990s) may contain information relating to one or both species (Clarke 1981;Jokinen 1982;Smith 2000), we have avoided literature that deals explicitly with C. japonica, and included literature pertaining to C. japonica only  Pace (1973), Köhler et al. (2012), Lu et al. (2014), and the Worldwide Mollusc Species Database (https://www.bagniliggia.it/WMSD/HtmSpecies/1071200630.htm). Question marks indicate areas where C. chinensis has been reported, but we consider indigenous status as uncertain or suspect. The confirmed range presented may be conservative, but it cannot be determined fully until taxonomy within the genus Cipangopaludina is resolved. This map was generated using ArcMap 10.7 with country polygon layers (ArcGIS Hub 2020) obtained from the DIVA-GIS repertoire (Diva-GIS 2020) and the world base map was sourced from ArcGIS Hub UIA World Countries Boundaries.
where genetic or morphologic distinction between the two is discussed (Burks et al. 2016;David and Cote 2019;Fox 2007;Hirano et al. 2015). Conversely, across eastern Asia, there are 16 species of Viviparidae, all relatively large. Hirano et al. (2015) identified three species of viviparids in Japan, Chiu et al. (2002) identified two in Taiwan, and Lu et al. (2014) identified 11 species and two subspecies within the genus Cipangopaludina in China. We also have avoided inclusion of those indigenous viviparids in our review, keeping the focus on C. chinensis in China, where it is abundant, and there is significant scientific literature on its indigenous habitats and ecological interactions.
Our list of references includes peer-reviewed literature, grey literature (e.g., reports, theses), and reputable internet sources. For this review, the term "publication" includes both peer-reviewed and grey literature sources. Together, these publications comprise the literature assessing C. chinensis effects on indigenous species in North America and Europe, the species establishment and spread within North America and Europe, and the history and biology of the species in both indigenous (Asian) and non-indigenous (North America and Europe) habitats. The European literature, however, is limited due to the species' very recent establishment on that continent ( Van den Neucker et al. 2017).
In regard to terminology, we use "non-indigenous" and "indigenous" as recommended by Colautti and MacIsaac (2004). Publications dealing with C. chinensis as an indigenous species in Asia were used to supplement the limited knowledge currently available for North America and Europe. Publications were identified as belonging to one of eight categories (Fig. 2) including (category names are noted in quotation marks, category names are listed in Supplementary Table S1 1 ): "Ecology" (ecological studies); "Reviews" (literature reviews); "Management" (species management); "Eco-system impacts" (disruption of systems by non-indigenous CMS); "Distribution" (both modelled output and absence/presence studies); "Commercial use" (use of CMS for industrial use and for profit); "Medical use" (use of CMS in medical studies and applications); and "Biology" (anatomy, taxonomy, life cycle, morphology, and genetic studies). Reputable internet sources were included in this review to determine the extent and types of information readily available to the public, but these sources were separated from publications because no formal research methodologies were followed or reported. Internet sources were considered "reputable" if they were managed by a group of researchers studying AIS, individual researchers, universities, or government organizations.

Conceptual diagrams
To predict and compare C. chinensis ecosystem interactions, conceptual diagrams presented here draw on literature dealing with C. chinensis across both its non-indigenous (Fig. 3A) and indigenous ( Fig. 3B) ecosystems using literature from the "Ecosystem impacts", "Ecology", and "Biology" categories, as well as suspected vectors of transport as a non-indigenous species in North America. The term "conceptual diagram" was selected because these diagrams provide a visual summary of the main conclusions derived from literature pertaining to C. chinensis biological needs, ecosystem impacts, and human interactions. These diagrams were created with two goals: (i) to determine if C. chinensis meets Canadian federal government criteria as a non-indigenous "invasive" species and (ii) to identify ecosystem interactions (either positive or negative) that may occur in North America should C. chinensis populations become as widely distributed or as abundant as the species is across its indigenous range. Literature located for this review (n = 93) dealing with Cipangopaludina chinensis and organized by nature of content. Publication sources were peer-reviewed journal articles (n = 86) and grey literature (n = 7). Additionally, source year of publication is plotted using a violin plot to visualize the relevance of the data. Most literature (Supplementary Table S1 1 ) were published 2010-2020, suggesting an increased interest in C. chinensis.

Geographical distribution of Cipangopaludina chinensis in Canada
We contacted multiple Canadian sources to gain access to and build a database of C. chinensis reported occurrences in Canada. An effort was made to include all Canadian reports of C. chinensis occurrence, regardless of year reported (i.e., both historical and more recent reports). Natural history collections databases from across Canada were accessed for records of C. chinensis. The following are the natural history museums contacted (and response to our query if they housed CMS reports): Royal British Columbia Museum (yes), Royal Alberta Museum (no), University of Saskatchewan Museum of Natural Sciences (no), Manitoba Museum (no), the Canadian Nature Museum (yes), the Royal Ontario Museum (yes), Ontario Natural Heritage Information Centre (yes), Quebec Biodome (no), New Brunswick Museum (yes), Nova Scotia Museum of Natural History (yes), and the Provincial Museum of Newfoundland and Labrador (yes). Additional reports were shared with us from the British Columbia Conservation Data Centre, Alberta Department of Environment and Parks, Environment and Climate Change Canada, Wildlife Systems Research, Fragile Heritage (Ontario), the Maritime Aboriginal People's Council (Nova Scotia, New Brunswick, and Prince Edward Island), and our own unpublished records (Kingsbury et al. 2021). The popular online citizen-science natural-history social networks, iNaturalist and Early Detection and Distribution Mapping System (EDDMapS) Ontario, both of which include an on-line app for smartphones, were also searched for confirmed observations of C. chinensis. All reports in our geographical distribution maps were verified from specimen vouchers, photographs, or from sources regarded as reliable, and geo-tags (latitude and longitude) confirmed. The geo-tagged report occurrences in Canada were compiled as a data layer in ArcGIS and compared with the USA C. chinensis data layer maintained by the USGS (https://nas.er.usgs. gov/viewer/omap.aspx?SpeciesID=1044). In keeping with our focus on Canada and continental North America, we also checked Mexican reports for C. chinensis. We did not find any "verified" species reports or reported presences in Mexican waterbodies. Cipangopaludina chinensis is not on the United Mexican States invasive species list or the Mexico Invasive Alien Species Action Plan (Mexico Secretariat of the Convention of Biological Diversity, n.d.).
The least common publication category was "Reviews" (two publications). These include a grey-literature review (Waltz 2008) and an older second review, which, although peer-reviewed, was written prior to the publication of the majority of the North American literature on C. chinensis (Jokinen 1982). Both reviews have been widely cited in a number of grey-literature documents, including federal government documents and theses (Bobeldyk 2009;Haak 2015;Rivera and Peters 2008), government and expert-run websites (Kipp et al. 2020), and peer-reviewed journal articles ( The 30 reputable internet sources found were from the USA (17), Canada (10), Australia (1), and international organizations such as the International Union for Conservation of Nature and Global Invasive Species Database (2020) (2) (see Supplementary  Table S2 1 for a full list of internet links). Much of the on-line information available to the public appears to be reputable (i.e., written by a scientific specialist) and is presented on government supported websites. These websites offer concise biological information, scientific and common names, descriptions of indigenous and non-indigenous range, and likely vectors of dispersal for C. chinensis in North America. Generally, internet sources tend to focus on the biology of the snail, often including notes on its indigenous range, and typically target recreational boaters in an effort to enlist the public in limiting the species introduction to water bodies where C. chinensis is non-indigenous and potentially invasive.

Life cycle
Literature from both indigenous and non-indigenous ecosystems indicate similar biological characteristics in all freshwater habitats, with a single C. chinensis female carrying embryos capable of rapidly establishing a population. Female C. chinensis give birth to live, shelled, young that are between 3-5 mm in shell diameter. Mature females can produce numerous offspring per brood (Fox 2007). Females reach sexual maturity at 6-12 months (Bobeldyk 2009;Zhang et al. 2017) and studies in North America estimate fecundity to be 27.2-33.3 young/female/year (Stephen et al. 2013). However, snail uteri have been documented holding as many as 102 embryos (Jokinen 1982;Waltz 2008). Chinese studies indicate that C. chinensis may reproduce 2-7 times/year (April-October) and brood 20-50 fertilized eggs at a time . In North America, females are believed to release juvenile snails from June to late October (Jokinen 1982). Haak (2015) found that adult C. chinensis exposed to colder temperatures (≤ 12°C) ceased reproduction, while those held at warmer temperatures (27°C) increased reproduction when compared to those at ambient temperatures (20°C). The number of brooded embryos appears to be directly correlated to female shell size in C. chinensis (Stephen et al. 2013). Cipangopaludina chinensis is believed to be iteroparous, with females expected to produce multiple broods over a lifetime (Fox 2007;Jokinen 1982), estimated at 5 years (Jokinen 1982;Stephen et al. 2013). Thus, once introduced to a site, C. chinensis may undergo exponential population growth, especially in the absence of density-dependent factors that might limit the population (Stephen et al. 2013).
Although habitat can influence the size and morphology of C. chinensis, this species is one of the largest freshwater snails among the Viviparidae (Liu et al. 1995). Growth of C. chinensis is indeterminate, with both sexes of the species reaching much larger maxima (40-60 mm) than any indigenous North America freshwater gastropod (Liu et al. 1995). Females appear to live longer than males (5 years versus 3-4 years) and therefore grow to larger size (Jokinen 1982). The larger shell size likely enables nonindigenous C. chinensis to avoid some predators that otherwise feed on aquatic North American gastropods (e.g., yellow perch, Perca flavescens) (Twardochleb and Olden 2016).

Morphology and genetics
Both morphology and genetics suggest that C. chinensis is a species distinct from C. japonica ( Fig. 4; Burks et al. 2016;Chiu et al. 2002;Clarke 1978Clarke , 1981Clench and Fuller 1965;David and Cote 2019;Hirano et al. 2015;Lu et al. 2014;Smith 2000;Wang et al. 2017). Nonetheless, the nomenclatural status of snails currently placed within the genus Cipangopaludina remains contentious (Clench and Fuller 1965;Smith 2000). This is especially the case across eastern Asia, where multiple subspecies of C. chinensis have been recognized and other species of Cipangopaludina co-occur (Lu et al. 2014). Lu et al. (2014) has develop an identification key for the 18 species of Cipangopaludina spp. recognized in China, which may be useful for tracking potentially invasive Cipangopaludina spp. in North America. Morphologically, two shell phenologies have been identified in C. chinensis; tall-spire and short-spire (Chiu et al. 2002). Short-spired shells are found in waterbodies with low pH and low calcium concentrations. Tall-spired shells are found in higher pH, higher calcium concentration lakes (Chiu et al. 2002). The differences in C. chinensis shell morphology further confounds identification issues between C. chinensis and C. japonica in North America, suggesting that one of the species may be more prevalent than previously believed. Recent research, using the mitochondrial cytochrome c oxidase subunit I (COI) gene, has established that C. chinensis is genetically distinct from C. japonica (Burks et al. 2016). To ensure correct species identification we recommend genetic testing, when possible, of mystery snails found in areas where both C. chinensis and C. japonica occur.

Indigenous habitat
Cipangopaludina chinensis (Chinese taxonomic name: 中国圆田螺) is wide-spread within its indigenous range and is an important prey option to indigenous predators and a staple-food item in the human diet in many regions of China and eastern Asia. The indigenous range of C. chinensis, commonly referred as the "mud snail" in China (Chinese common name: 田螺) extends across eastern Asia, although the true extent of range remains unknown due to taxonomic uncertainty (Fig. 1). Cipangopaludina chinensis occurs naturally in rice paddies, rivers, and lakes, where it is common and where positive ecosystem interactions have been documented (Dewi et al. 2017;Kurniawan et al. 2018;Liu et al. 1995;Nakanishi et al. 2014). As a result, this species is an important component of managed rice paddy ecosystems where the presence of mud snails is correlated with a greater abundance and diversity of terrestrial arthropods, which leads to greater rice yields (Dewi et al. 2017). Also, within its indigenous range, C. chinensis is prey for multiple species (Supplementary Table S3 1 ), including waterfowl such as Muscovy duck (Cairina moschata, 番鸭), domestic duck (Anas platyrhynchos domesticus, 北京鸭), and mallard duck (Anas platyrhynchos, 绿头鸭), Asian carp, such as black carp (Mylopharyngodon piceus, 青鱼) and Amur carp (Cyprinus rubrofuscus, 鲤鱼), and Chinese softshell turtle (Pelodiscus sinensis, 中华鳖). The species is also sold in local wet markets for human consumption (Luo et al. 2012;Tian et al. 2012;Yan 2002;Zhang et al. 2017). Cipangopaludina chinensis also competes ecologically with the golden apple snail (Pomacea canaliculata, 福寿螺), which is a non-indigenous invasive species in China (Luo et al. 2012;Zhang et al. 2017). In North America, species reported preying on C. chinensis include largemouth bass (Micropterus salmoides), pumpkinseed sunfish (Lepomis gibbosus), signal crayfish (Pacifastacus leniusculus) Twardochleb and Olden 2016), and ringed crayfish (Faxonius neglectus) (Kelley 2016). As predators, these species, and others as yet identified or investigated (e.g., freshwater drum Aplodinotus grunniens), warrant further examination as potential biological controls of non-indigenous C. chinensis.

Ecological and biological tolerances
Due to its ecological plasticity, C. chinensis has proven to be extremely adaptable to a variety of environments (Table 1). Burnett et al. (2018) tested thermal tolerance in C. chinensis, establishing temperature tolerances ranging from 0 to 45°C, with 0°C the lowest temperature tested (and not necessarily the lower thermal limit). These experiments assessed critical maximum temperature and incipient lower lethal temperature over acute exposures (i.e., relatively short-term exposure). Cipangopaludina chinensis survival decreases where oxygen concentration is below 15 mg/L, and 4-day survival rate drops to 38.3% at pH 5.5, suggesting the species is Fig. 4. The three mystery snail species, Chinese (Cipangopaludina chinensis), Japanese (C. japonica), and banded (Viviparus georgianus), established in continental North America. The shell of C. japonica (6-7 whorls, %65 mm shell length, 65-80°spire angle) is reported to have a taller spire, a more acute spire angle and finer carina (arrow) on the shell surface than compared to C. chinensis (7-8 whorls, %50-65 mm shell length, 50-55°spire angle). Cipangopaludina chinensis is broader than C. japonica. However, these distinguishing features appear to be plastic and shell morphology may vary considerably. Molecular genetic methods are the most reliable method of distinguishing C. chinensis from C. japonica. Viviparus georgianus is the most distinguishable of the three, with a yellow-green shell, dark brownish-red bands, 4-5 whorls, %45 mm shell length, and 70-85°spire angle. Catalogue numbers of illustrated specimens: CM58-143 (C. chinensis), CM47465 (C. japonica), and CM62.6630 (V. georgianus). Photo by Dr. T.A. Pearce, Section of Mollusks, Carnegie Museum of Natural History.  Jokinen 1982 Environmental magnesium range at sites of species occurrence; no laboratory experiments were conducted to establish these thresholds. Sodium concentration 2-49 ppm Jokinen 1982 Environmental salinity range at sites of species occurrence; no laboratory experiments were conducted to establish these thresholds.
sensitive to oxygen and pH levels . Molluscicides, such as copper sulphate and rotenone, are not effective at controlling this species (Haak et al. 2014). Although an aquatic mollusc, initial air exposure experiments found that adult snails can survive at least 4 weeks of air exposure and juvenile snails survive 3-14 days (Havel 2011). Other desiccation experiments suggest that adult C. chinensis can survive exposure for longer than 9 weeks (Unstad et al. 2013). Finally, the concentration of calcium required (Table 1) by C. chinensis, at < 2 ppm, is very low compared to other molluscs, such as the zebra mussel (Dreissena polymorpha), quagga mussel (Dreissena rostriformis bugensis), and other calcifying AISs such as banded mystery snail and rusty crayfish (Faxonius rusticus) (Chiu et al. 2002;Latzka et al. 2015). Our conceptual diagrams therefore represent a conservative estimate of C. chinensis ecosystem impacts in North America. Notably, the ecological and biological data available (Table 1) is often based on opportunistic accounting of where the species has been discovered and does not reflect statistically robust ecological threshold experiments. Therefore, further research is needed to determine the ecological thresholds of this species.

Parasitology
In North America it appears that parasite prevalence in C. chinensis is much lower than among indigenous gastropod species (Karatayev et al. 2012). Of 147 necropsied C. chinensis from lakes in Wisconsin, USA, only two contained trematode parasites (Harried et al. 2015).
One snail contained Cyathocotyle bushiensis, a parasite of waterfowl known to cause mortality in dabbling ducks (Hoeve and Scott 1988), and the other contained nine aspidogastrean flatworms, Aspindogaster conchicola (Harried et al. 2015), a widespread parasite (North American, Europe, Africa) of freshwater bivalves, fish, and turtles. Karatayev et al. (2012) found that of 30 C. chinensis collected from the Great Lakes area, Canada, none were infected with parasitic helminths. In laboratory experiments where C. chinensis and two indigenous snail species, Bithynia tentaculata and Physa gyrina, were exposed to Sphaeridiotrema pseudoglobulus, an important waterfowl parasite, C. chinensis contained significantly fewer metacercariae than either of the two indigenous snail species (Harried et al. 2015). Moreover, those metacercarial cysts found in C. chinensis were typically encased within the mollusc shell and likely no longer viable (Harried et al. 2015). This limited information suggests that C. chinensis may be less vulnerable to infection by endoparasites commonly affecting indigenous gastropod species in North America. However, it is concerning that C. chinensis has been found hosting A. conchicola as this parasite has been linked to decreased reproduction in North American freshwater mussels (Gangloff et al. 2008), a diverse group of organisms for which there is considerable conservation concern (Williams et al. 1993).
East Asian research demonstrates that C. chinensis may act as an intermediate host for a number of human endoparasites of medical importance, including Echinostoma cinetorchis (an Asian intestinal Note: Many of the references report ecological data for locations where the species has been discovered and are not actual ecological threshold experiments, and as such should be interpreted carefully. There is a need for laboratory studies to assess the true ecological thresholds for this species in order to assess potential habitats (see Kingsbury et al. 2021). Literature references with an asterisk (*) denote references from C. chinensis indigenous range (i.e., Eastern Asia). fluke) and Angiostrongylus cantonensis (a nervous system nematode) (Chao et al. 1993;Chung and Jung 1999;Jokinen 1982;L€ u et al.2006;Sohn and Na 2017). Although of Asian origin, A. cantonensis has been identified as an emerging zoonotic pathogen in North America (York et al. 2015).

Biotransfer of contaminants
There have yet to be studies published on the potential for C. chinensis in North America to transfer contaminants from environments with elevated contaminant concentrations to human consumers. Nonetheless, several North American studies have shown that this species bioaccumulates contaminants present in polluted sediments, specifically mercury, arsenic, iron, manganese, zinc, copper, nickel, lead, cadmium, and chromium (Chapman et al. 2020;Tornimbeni et al. 2013). The study of the biotransmission of contaminants from sediments to C. chinensis and thence to consumers has been a key focus of multiple studies in Asia (Cui et al. 2012;Fang et al. 2001;Luo et al. 2016;Wu et al. 2001). Mainly, this is because of human health concerns due to the wide-spread consumption of C. chinensis in Asia. For example, of 14 edible mollusc species purchased at food markets originating in the Pearl River Delta, China, C. chinensis was among those with concentrations of cadmium, copper, zinc, lead, nickel, chromium, antimony, and tin that exceeded the regulatory human consumption limits. Chromium, most notably, exceeded the Chinese daily consumption limits of 1 lg/g (Fang et al. 2001). Snails from the highly polluted Zhalong Wetland, China, also had elevated lead (average concentration = 45.32 lg/kg), cadmium (average concentration = 2.45 lg/kg), and arsenic (average concentration = 11.48 lg/kg), due to grazing over highly contaminated sediments. (Luo et al. 2016). Additionally, C. chinensis has been documented to bioaccumulate zinc and copper from sewage sludge used to fertilize rice paddy fields . While the ability to bioaccumulate contaminants may prove useful in bioremediation, or allow C. chinensis to serve as a bioindicator, there is also concern that where the species is a regular part of the human diet, that C. chinensis may transfer contaminants to consumers. Even though C. chinensis is consumed far less frequently by humans in North America than in Asian countries, further North American research on bioaccumulation and contaminant transfer is required. Some North American wildlife certainly consume C. chinensis and may bioaccumulate contaminants that may impair North American wildlife populations or eventually find their way to human consumers.

Inter-mollusc competition
Where C. chinensis is present in North America, especially where other freshwater invasive calcified species are established (i.e., gastropods, bivalves, crayfish), there is evidence of impacts on indigenous species. Studies by Sura and Mahon (2011) found that the presence of C. chinensis caused snails (Helisoma trivolvis) to increase feeding rates and suggested cascading changes to community structures due to higher rates of algal consumption by C. chinensis (Sura and Mahon 2011). Johnson et al. (2009) found that indigenous snails (H. trivolvis, Lymnaea stagnalis, and Physa gyrina) decreased in mass during in-laboratory exposure to C. chinensis. In the presence of a second invasive species, the rusty crayfish, one indigenous mollusc species became extinct (Johnson et al. 2009). The thick shell and larger size of C. chinensis relative to indigenous gastropods leaves C. chinensis more resistant to predation by rusty crayfish (Johnson et al. 2009). However, rusty crayfish and other invasive crayfish species, such as red swamp crayfish (Procambarus clarkii) and northern crayfish (Faxonius virilis), will consume C. chinensis, although apparently taking only smaller, juvenile snails. Indigenous signal crayfish seem to be particularly prone to prey on C. chinensis ). Predation experienced by C. chinensis in mesocosm experiments with red swamp, virile, and signal crayfish decreased with increased C. chinensis shell size ). As a result, C. chinensis has been documented to have an impact on inter-species dynamics in regard to competition, stressors, and as a potential prey source for a limited number of indigenous predator species.
There is evidence that the diet, feeding mechanisms, and excretions of C. chinensis may alter freshwater bacterial and algal communities leading to changes in water chemistry and food webs (Bobeldyk 2009;Chen et al. 2011;Johnson et al. 2009;Olden et al. 2013). Previous studies have shown that C. chinensis ingest benthic organic matter, inorganic matter, and algae, primarily using the radula. Stomach contents of C. chinensis show that this species favours diatoms, but other studies indicate that C. chinensis does not feed selectively (Jokinen 1982;Olden et al. 2013;Plinski et al. 1978;Sta nczykowska et al. 1971). Additionally, larger adult C. chinensis (≥ 44 mm) can filter feed and will do so selectively when inter-snail competition is high (Olden et al. 2013). The ability to switch between two feeding mechanisms has previously been noted as a competitive advantage (Brendelberger and Jurgens 1993). Freshwater gastropods are typically grazers and freshwater bivalves are filter feeders (Brendelberger and Jurgens 1993). However, the ability of C. chinensis to compete for nutrition via both feeding mechanisms gives it a competitive advantage. Also, due to the significant amounts of nitrogen and phosphorous this large gastropod excretes via fecal matter, it is possible for C. chinensis to alter the algal community in habitats occupied (Olden et al. 2013). Mesocosm experiments with high C. chinensis densities showed significant changes in bacterial community composition (Olden et al. 2013). Changes in bacterial community composition led to a decrease in bacterial community variability, but did not decrease bacterial abundance (Olden et al. 2013).

Nutrient cycling
Filtration rates of large C. chinensis (maximum 471 mL/snail/h) are comparable to high-profile invasive freshwater and marine bivalves, including zebra mussel, quagga mussel, Asian clam (Corbicula fluminea), golden mussel (Limnoperna fortunei), and blue mussel (Mytilus edulis) (Olden et al. 2013). Additionally, mesocosm studies with C. chinensis present at low and high densities showed decrease of 54% in the amount of suspended chlorophyll-a concentrations at "low" snail density and a 115% decrease at "high" snail density (Olden et al. 2013). Substrate chlorophyll-a increases observed were likely due to elevated nitrogen:phosphorus (N:P) concentrations ratios produced by C. chinensis excreta that promote periphyton production (Johnson et al. 2009;Olden et al. 2013). Previous research suggests that C. chinensis increase N:P ratio in the water column due to low P excretion (Johnson et al. 2009). Future research should further investigate the food web implications of bacterial community shifts and the species of algae produced (whether C. chinensis presence may lead to higher probabilities of toxic algal blooms). Previous studies link C. chinensis nutrient cycling, particularly nitrogen (N) and phosphorus (P) compounds to downstream eutrophication, with higher temperatures producing greater release rates of total-N, dissolved-N, total-P, dissolved-P, ammonia, and phosphate (Bobeldyk 2009;Chen et al. 2011). These studies suggest that C. chinensis may play an important role in eutrophication. This warrants further research to determine C. chinensis population densities required to cause eutrophication. More importantly, aquatic systems already under pressure of elevated nutrient levels, such as from farmland run-off, should be protected against C. chinensis introduction to ensure that this species does not become the tipping point that leads to eutrophication of these vulnerable ecosystems.

Commercial use
An array of human uses for C. chinensis have been documented in East Asia, although in North America C. chinensis has not yet been exploited commercially. The species has been used as feed for larval fireflies (Luciola ficta) (Ho et al. 2010) and hatcherycultured mud crab (Scylla paramamosain) (Gong et al. 2017). Fish farms in China use viviparid snails as commercial feed for Chinese carp (M. piceus and Cyprinus rubrofuscus) (Yan 2002), Chinese softshell turtles (Tian et al. 2012), and Chinese golden-coin turtles (Cuora trifasciata) (Luo et al. 2012). Cipangopaludina chinensis has been used to remediate rice paddy fields contaminated with sewage with elevated metals (Fujibayashi et al. 2016;Xing et al. 2016).

Medicinal use
In Asia, there have been recent efforts to develop pharmaceutical uses for C. chinensis. Cipangopaludina chinensis has been long used in traditional Korean medicine and is a component of traditional Korean medical knowledge used to treat indigestion (Kim et al. 2018). More recent medical studies have explored the efficacy of compounds derived from various C. chinensis tissues for protection against liver damage Wang et al. 2015a;Xiong et al. 2019;Yang et al. 2012), as an anti-inflammatory for joint pain-caused immune responses (Lee et al. 1998;Maoka et al. 2012;Shi et al. 2016), and as a supplement that protects against plaque build-up in arteries (Xiong et al. , 2017(Xiong et al. , 2019. Additionally, C. chinensis has been used as a model organism in neurophysiological, hepatological, and reproductive studies testing pharmaceutical products and supplements (Swart et al. 2017;Wang et al. 2014;Wang et al. 2015). Finally, waste shells have been recycled to extract calcium that can be used for promoting bone growth (Zhou et al. 2016). These methodologies have the potential for transfer to, and development in, North America as a means of reducing non-indigenous C. chinensis populations.

Management
Commonly employed molluscicides, including rotenone and copper sulphate, are not effective in controlling or eliminating C. chinensis (Haak et al. 2014) and can be toxic to other invertebrates. Drawdowns, in which the water is removed from a specific water body, will likely not eradicate C. chinensis because the species can survive periods of desiccation exceeding 9 weeks (Unstad et al. 2013). Furthermore, long periods of air exposure, or increased predator pressure, has only been found to elevate C. chinensis reproductive rates (Prezant et al. 2006;Unstad et al. 2013; and personal observation of laboratory cultures). Juvenile C. chinensis typically hide under rocks and burrow into sediment, so it is impractical to manually cull populations of C. chinensis (Jokinen 1982).
Nonetheless, approaches to management, either through eradication and (or) limiting further introductions, have not been thoroughly explored, with only five publications available Haak et al. 2014;Matthews et al. 2017;Rothlisberger et al. 2010;Unstad et al. 2013). Only two of these publications conducted experiments to examine C. chinensis response to commonly used AIS management approaches (Haak et al. 2014;Unstad et al. 2013). Two publications have focused on public education, placing controls on species movements, and early species detection Rothlisberger et al. 2010). Matthews et al. (2017) noted that there are no ongoing case studies for eradication of C. chinensis, perhaps due to the perception that C. chinensis in North America is always present at relatively low densities and is nonproblematic. Currently, species monitoring of C. chinensis in North America is lacking in terms of monitoring population density. Population density likely plays an important role in how C. chinensis impacts indigenous species. As previously noted, C. chinensis at relatively low densities is able to overwhelm nutrient release of indigenous molluscs (Bobeldyk 2009). Eradication of AIS (by chemical, manual, biological, or mechanical means) is generally only effective when a species is recently introduced to an aquatic system (Government of Canada 2019a), supporting the need for an effective monitoring and modelling program across Canada. A single study of C. chinensis ecological preference for slow moving water noted that perhaps increasing water flow via installation of a culvert between connected water bodies may limit the upstream spread of C. chinensis. However, these results were only observational and require further investigation (Rivera and Peters 2008). As we note elsewhere in this review, C. chinensis has been documented as established in a variety of freshwater, and even brackish habitats (see section titled "Geographical distribution of reported C. chinensis occurrences in Canada"), and under various flow regimes.
Some Canadian government agencies have taken steps to limit the spread of C. chinensis by preventing new introductions. In Canada, Manitoba and Alberta both have laws controlling the sale, distribution, and reporting of C. chinensis and other nonindigenous viviparid snail species. These Canadian provinces have also implemented required boat cleaning and AIS reporting programs to ensure AIS are not accidently transferred between waterbodies and to improve early AIS detection (Government of Alberta 2015; Government of Manitoba 2015). The DFO, with federal oversight of AIS in Canada, operates under the Fisheries Act and enforces AIS management through the Aquatic Invasive Species Regulations. Section 10 of the Aquatic Invasive Species Regulations prohibits any person from introducing a species that is non-indigenous to any region or water body frequented by fish unless authorized under provincial or federal law (Canada Legislative Services Branch 2015). Cipangopaludina chinensis is listed under the Aquatic Invasive Species Regulations, and is therefore subject to management under these regulations. However, a previous risk assessment by DFO has assessed C. chinensis impacts to be negligible in Canada and this has led to reduced management priority for C. chinensis when compared to other AISs (Schroeder et al. 2013). In the USA, examples of governmental control of C. chinensis include Missouri and Minnesota. Cipangopaludina chinensis is listed as a "regulated invasive species" in Minnesota, which allows for possession, sale, purchase, and transportation of this species, but prohibits the introduction of live C. chinensis into new ecosystems, such as through aquarium releases or via garden ponds (Minnesota Department of Natural Resources 2020). In Missouri, C. chinensis is designated as a "prohibited species", meaning that no person may possess, purchase, sell, transport, import, or export this species (Missouri Department of Conservation 2020). Within the European Union a risk assessment was completed that recommended targeting the pet and aquarium trade and increasing public education/engagement to assist in controlling the further spread of C. chinensis . It is clear that regulatory action to prevent the sales or transport of viviparid snail species via the pet, aquarium, and water garden trade, and public education and engagement will be essential for early detection and slowing the spread of C. chinensis in North America.
In Canada, public education concerned with invasive and potentially invasive species, including C. chinensis, is limited ; Office of the Auditor General of Canada 2019). This has led to a lack of public engagement in terms of monitoring invasive species across much of Canada, with the possible exception of Ontario. This may explain some of the gaps in distribution currently present in Canada. A more complete understanding of C. chinensis' impacts on freshwater ecosystems in North America is required, especially in terms of determining where C. chinensis is present, and which ecosystems are likely to be most affected. Also, species managers should consider expanding the terminology used to describe C. chinensis from merely "non-indigenous" or "non-native" to incorporate "invasive". This should encourage greater public support for, and assistance with, species containment and geographical analyses.

Conceptual diagrams
Our conceptual diagram of C. chinensis ecosystem interactions in North America (and Europe) further supports the argument that Material data 1 and statistically analysed these data to determine if each of these parameters impacted the density of C. chinensis found at each site. We did not find any statistically significant correlations when the Otter Lake outlier was excluded.

Shoreline development
Greater shoreline development = greater snail density 1.00 6 0.40 (p = 0.005) Proximity to population centre Closer proximity = greater density -0.71 6 0.37 (p = 0.04) Conductivity Secchi disc depth -0.84 6 0.4 (p = 0.02) Boat lunch presence Positive correlation C. chinensis is an AIS on the continent (Fig. 3A). Cipangopaludina chinensis negatively affects indigenous aquatic species by increasing environmental pressures (e.g., greater resource competition, changes in food web structure, changes to nutrient cycling), leading to decreased biomass and even localized extinction of indigenous molluscs (Johnson et al. 2009;Karatayev et al. 2009;Sura and Mahon 2011). Indirect, negative, impacts on indigenous aquatic species may be caused by C. chinensis through shifts in food web structure (Olden et al. 2013), changes in water chemistry (Olden et al. 2013), and eutrophication (Bobeldyk 2009). Also, C. chinensis has become a vector for indigenous A. conchicola, which parasitize North American unionid bivalves of conservation concern (Harried et al. 2015).
The indigenous (Asian range) conceptual diagram identifies areas where North American literature is incomplete and flags some future concerns for growing C. chinensis populations and further distribution in North America (Fig. 3B). In particular, consumption of C. chinensis by North American wildlife species, and perhaps humans, is potentially problematic because there is good evidence of C. chinensis transferring parasites and contaminants to both human and wildlife consumers in Asia (Chao et al. 1993;Chung and Jung 1999;Cui et al. 2012;Fang et al. 2001;Lu et al. 2018;Luo et al. 2016;Sohn and Na 2017;Tornimbeni et al. 2013;Wu et al. 2001). Further research is required to assess the importance of C. chinensis in hosting parasites or transferring contaminants into North American wildlife consumers that are part of freshwater food webs, or to humans.
Additionally, our literature review and conceptual diagrams uncovered gaps in current knowledge for C. chinensis distribution and management across Canada. There are indications that C. chinensis is an invasive species with the potential to disrupt North American ecosystems and food webs and to reduce the economic benefits of freshwater habitats (Fig. 3A). Unfortunately, available data are based on a limited number of laboratory and field experiments and surveys (Bury et al. 2007;Chaine et al. 2012;Clarke 1981;Haak et al. 2017;Karatayev et al. 2009;Kipp et al. 2020;Latzka et al. 2015;McAlpine et al. 2016;Rothlisberger et al. 2010;Solomon et al. 2010;Stephen 2017). Also, there is limited literature regarding C. chinensis in its indigenous ranges that can be used to supplement the English-language literature and support management decisions. While, in theory, formal risk assessments may allow one to accurately assess the risks associated with AIS in North America, data and studies are still needed to support any such risk assessments for C. chinensis. A systematic, continental-scale, record-keeping approach to C. chinensis in North America would better enable predictions of habitat overlap with indigenous species of conservation concern (e.g., species at risk).
The conceptual diagrams indicate that C. chinensis' role in North American habitats (Fig. 3A) versus its role in Asian habitats ( Fig. 3B) are similar. However, the outcomes of interactions between C. chinensis and its indigenous ecosystems are less extreme in the species Asian range than in North America. Presumably, this is evidence of a current lack of ecological equilibrium in North America between C. chinensis and the ecosystems it occupies on the continent. It is also worth noting that research priorities involving C. chinensis in Asia and North America differ and that the conceptual diagrams presented here are visual summaries of current available literature. In North America (and Europe), research has typically focused on determining the ability of C. chinensis to disperse and become established, and whether the species exhibits invasive behaviours. In Asia, work on C. chinensis has often been concerned with issues of environmental contamination and monitoring and possible commercial or medicinal use of C. chinensis Li 2012). In summary, the introduction of a non-indigenous species into an ecosystem can have unpredictable outcomes. It is likely that the ecosystem impacts C. chinensis exerts on invaded habitats may be density dependent. Currently, no studies exist that compare snail density in its native range versus the non-indigenous range. However, a few studies (Table 2) do suggest certain environmental parameters that may impact C. chinensis successful establishment and population growth (Chaine et al. 2012;Collas et al. 2017;Haak et al. 2017;McCann 2014;Nakanishi et al. 2014;Solomon et al. 2010). Our "non-indigenous (North America and Europe)" and "indigenous (Asia)" conceptual diagrams should be useful for assessing potential scenarios, in they summarize previously published information on C. chinensis from both indigenous and non-indigenous ranges. Based on the conceptual diagram presented here and applying the DFO definition for an AIS (Fisheries and Oceans Canada 2018), C. chinensis in North America should be regarded as an invasive snail.

Geographical distribution of reported C. chinensis occurrences in Canada
We located 278 reports documenting C. chinensis in Canada (Fig. 5), with the greatest number of reports from Ontario (77.7%). Canadian reports of C. chinensis occurrence by province included: British Columbia (12); Alberta (one); Ontario (216); Quebec (10); New Brunswick (seven); Nova Scotia (30); Prince Edward Island (one); and the island of Newfoundland (one). We located no reports of C. chinensis from Saskatchewan, Manitoba, Yukon, Northwest Territories, Nunavut or Labrador (Table 3). The first reported occurrence in Canada was at St. John's, Newfoundland, in 1905, the second from Kawkawa Lake, British Columbia (1953. Since then there has been only sporadic reporting in Canada (Clarke 1981;Clench and Fuller 1965;Kalas et al. 1980;McAlpine et al. 2016; Plinski  . 1978;Sta nczykowska et al. 1971). The greatest number of Canadian C. chinensis occurrence reports (Fig. 6) were made between 2010-2019 (215 reports), 1990-1999 (33), followed by 2000-2009 (26). Very few reports pre-date 1990 (nine). For example, C. chinensis was first noted in Quebec in 1971 (Jokinen 1982), but there have been very few subsequent reports for the province. Also, in Ontario, where monitoring and reporting of C. chinensis occurrences has been most consistent, reports indicate that the species is spreading quickly to new water bodies from invaded ones (Fig. 6). This is concerning due to the species wide range of ecological tolerances and suggests that this species is likely to continue spreading.
The source of C. chinensis reports varied and is associated with the level of public awareness in each province, and whether there is an easily accessible means of reporting. In British Columbia, C. chinensis reports were derived entirely from museum records.
The Prairie provinces (Alberta, Saskatchewan, Manitoba) had no museum records or published reports at the outset of preparing this review. However, Alberta Environment recently confirmed C. chinensis as well established in Lake McGregor in southern Alberta, and the provinces has now started an awareness campaign (Alberta Invasive Species Council 2018). This suggests that all Prairie provinces may need to add this species to their nonindigenous monitoring programs. Ontario reporting methods include a preponderance of internet and phone app reports, likely the result of a high level of public awareness of invasive species through educational campaigns. Publicly accessible phone apps and internet databases, such as the Ontario Department of Natural Resources online database (EDDMaps Ontario; 33% of our reports) and iNaturalist (21% of our reports), were the most frequently used tools for reporting public sightings of C. chinensis in  Kipp et al. (2020). Hawaii is excluded, but it is the site of the first reported case of C. chinensis in North America (year 1800). There are four reports from the Hawaiian Islands (Hawaii, Kauai, Molokai, and Oahu), but none since 1997 (Kipp et al. 2020). Canadian records (red) were assembled for this study (see text for sources). The gaps in species occurrence reports suggest investigator bias (i.e., only reported by individuals with knowledge of this species and access to reporting resources such as phone apps). Therefore, this map is likely a conservative estimate of the species' true distribution. Species occurrence points are aggregated to within 10 km. This map was generated using ArcMap 10.7 with base maps sourced from Statistics Canada (2019) (Canada base map, rivers, lakes, and coastal waters) and ArcGIS (2012, 2013) (USA and Mexico base maps). This is the first collected and mapped national database of C. chinensis occurrences in Canada (see subsection titled "Geographical distribution of C. chinensis in Canada" in section "Approach" section for details) and the USA point data were sourced from United States Geographical Services (Kipp et al. 2020).
Ontario. In Quebec, the majority of C. chinensis reports were derived from previously published literature (71.4%) with a limited number of reports available at iNaturalist (28.6%). In Atlantic Canada (New Brunswick, Nova Scotia, Prince Edward Island, Newfoundland, and Labrador) scientists researching freshwater lakes (50%) were the most common source of C. chinensis records, followed by museum records (25%), and iNaturalist (11%).
Across Canada, established populations of C. chinensis have been reported from 146 locations (Table 3): lakes (64), rivers (29), and ponds (25) were the most frequently reported habitat of C. chinensis. However, there were some reports of C. chinensis occurrence in marshes (1), bays (5), and harbours (1). The number of freshwater systems with reported C. chinensis occurrences by province include British Columbia (11), Alberta (1), Ontario (105), Quebec (9), New Brunswick (6), Nova Scotia (16), Prince Edward Island (1), the island of Newfoundland (1), (Table 3). The number of invaded water bodies per province followed the same trend as the number of reported occurrences of C. chinensis with the highest number of invaded water bodies in Ontario (Table 3). Previous studies have suggested that the habitat preference of C. chinensis as slow or stagnant freshwater McAlpine et al. 2016;Rothlisberger et al. 2010). Our review of Canadian C. chinensis reports indicate that C. chinensis also inhabit rivers, harbours, and bays, which agrees with Asian literature on this species use of habitat across its indigenous range. Assumptions concerning C. chinensis habitats in North America can be risky. For example, in 2019, C. chinensis was confirmed as present in a brackish tidal stream (conductivity = 349.8 ls/cm) directly connected with Cow Bay (conductivity = 36 700 ls/cm) in the Halifax Regional Municipality (Fraser, personal observation 2019). This leads to questions about salinity tolerance and thresholds for this species in locations where seasonal salt and freshwater mixing may occur. This is the first country-wide summary for C. chinensis in Canada. A federally-maintained Canadian database would allow researchers and managers to identify what will likely be an expanding distribution, gaps in distributional survey effort, and most importantly, will contribute to predictive distribution models (Haak et al. 2017;Papes et al. 2016;Kingsbury et al. 2021) that may highlight areas of concern (e.g., species and habitats at risk, areas of ecotourism, recreational fishing spots). Species distribution modelling for C. chinensis in Atlantic Canada and other parts of North America suggest that water chemistry and ecosystem composition are important factors for successful C. chinensis establishment, and human-mediated transfers between waterbodies are significantly correlated with C. chinensis introductions (Haak 2015;Haak et al. 2017;Kingsbury et al. 2021;Latzka et al. 2015;Papes et al. 2016). Environmental parameters ( Table 1) that may impact C. chinensis establishment include calcium concentration, flow rate, pH, maximum water depth, and distance to highway (Chiu et al. 2002;Collas et al. 2017;Fraser et al., unpublished;Haak 2015;Jokinen 1982;Latzka et al. 2015), and there may be others as yet unidentified. Studies focused on factors influencing successful establishment of C. chinensis have been to date limited in North America, both in terms of geographic area and in modeling success. In particular, a lack of knowledge of ecological thresholds and ecosystem interactions between C. chinensis and other indigenous or invasive species, inconsistency in water chemistry databases, and inconsistent coverage of water chemistry databases across political jurisdictions (e.g., a whole province or state) has limited research outcomes. Research and improved monitoring efforts, both for species occurrence and habitat characteristics, are clearly needed to expand our current understanding of how ecological thresholds influence habitat selection in C. chinensis, as these will influence presence/absence surveys and predictive distributional modeling for the species.
Not surprisingly, current Canadian C. chinensis distribution is concentrated in areas where human population is most dense. This may reflect actual distribution, but it may also be reflective of reporting opportunity. The distribution of C. chinensis in Canada is essentially a northward extension of the species range in the USA, so spatial trends for this species south of the Canadian border merit consideration when examining Canadian trends. The geographical distribution of North American C. chinensis reports in both Canada and USA identified a few likely reporting gaps in western Canada, around Lake Superior in Ontario, southern Quebec, and eastern Canada (Fig. 5). For example, C. chinensis was first noted in Quebec in 1971 (Jokinen 1982); but there have been very few subsequent reports for the province, even though the snail is commonly found in some areas of the Ottawa and St. Lawrence rivers (A. Ricciardi, McGill University, personal communication, 2020). The prevalence of pleasure craft, developments that redirect and change water flow, and the significant likelihood for aquarium releases suggest a wider establishment for C. chinensis in southern Quebec since 1971 . Public education of C. chinensis as a potential problem for indigenous ecosystems, in parallel with an accessible reporting system (e.g., smartphone apps, email, mobile texting, and phone contacts), should lead to increased reports of occurrence. We recommend that a Canada-wide database of C. chinensis occurrences, using the compilation we present here as a foundation, be established, and perhaps maintained by DFO. These occurrence reports can then be included with other geo-tagged data layers for non-indigenous species, vulnerable species and species-atrisk, and vulnerable habitats to determine areas of conservation management priority.

Conclusion
We reviewed 123 literature and internet sources (86 peer-reviewed journal articles, seven grey literature, and 30 internet sources) for both indigenous and non-indigenous C. chinensis habitats, compiled North American reported occurrences, and created two conceptual diagrams for both indigenous (Asia) and non-indigenous habitats (North America and Europe). Evidence suggests that C. chinensis meets the current federal Canadian definition of an aquatic invasive species.
While we believe the geographical distribution of reported occurrences that we present under-represents actual C. chinensis distribution in Canada, this is the first Canada-wide compilation. The data indicate that C. chinensis is already widespread in Canada. Our conceptual diagrams suggest freshwater ecosystems in North America may be vulnerable to the negative impacts of C. chinensis. Our indigenous (Asian range) conceptual diagram suggests consumption of C. chinensis by North American wildlife species or humans is problematic, as there is good evidence of C. chinensis transferring parasites and contaminants to both humans and wildlife consumers in Asia. With global temperatures predicted to rise, and given the species adaptability and fecundity, C. chinensis can be expected to expand its range northward in North America. Additionally, rising temperatures are predicted to lead to an increase in rates of release of nitrogen and phosphorous compounds, impacting food webs by altering bacterial community structure and potentially leading to an increased risk of eutrophication where C. chinensis becomes established and abundant. Due to the difficulty of eradicating C. chinensis once established, preventing introduction of the species is the most cost-effective management strategy. This will require increased public education and improved regulations that includes the enforcement of existing AIS regulations on the sale of C. chinensis in the aquarium and water garden trade and prohibition on the release of the species into the environment in Canada. Canada lags in terms of research on C. chinensis, with most of the northern temperate literature based on studies from the USA and Europe. While further research is needed in North America to better document C. chinensis distribution and to establish predicted impacts, impacts may vary geographically. Those in Canada may depart from impacts documented to date in the southern USA. In summary, we recommend that C. chinensis be recognized as an aquatic invasive species (AIS) in Canada (and across the USA) and that the species be managed through a combination of public education and enforcement of regulations.

Acknowledgements
Funding for this research was provided by Saint Mary's University, Fisheries and Oceans Canada, Veterans Affairs Canada, CLARI Fig. 6. The number of Cipangopaludina chinensis occurrences reported at 5-year intervals across Canada (plot on the left) follows a similar trend to the number of water bodies reported invaded by C. chinensis each year in Canada in each province (plot on the right). The frequency of reports made may be increasing due to advancements in technology (e.g., smart phone apps), which may also impact the number of water bodies reported invaded each year. Regardless, the number of water bodies invaded by C. chinensis in Canada is increasing. AB, Alberta; BC, British Columbia; NB, New Brunswick; NFL, Newfoundland and Labrador; NS, Nova Scotia; ON, Ontario; PEI, Prince Edward Island; QC, Quebec.
Case Fund, and the New Brunswick Museum Florence M. Christie Scholarship. We would also like to thank the members of the Dynamic Environments & Ecosystem Health Research group at Saint Mary's University for their help in collecting C. chinensis reports, their assistance in lake surveys, and their numerous edits. Special thanks to Meghan Fraser and Megan Fong for assisting with creation of the Canada map, C. chinensis culture maintenance, and considerable time dedicated to the Chinese Mystery Snail Project. Lastly, thanks to Timothy Pearce and the Carnegie Museum of Natural History for providing the photograph of the non-indigenous mystery snails found in North America. We greatly appreciate access to the Carnegie Museum of Natural History mollusc collection and associated expertise.