Aquatic Invasions (2019) Volume 14, Issue 2: 267–279

CORRECTED PROOF

Research Article A rare example of non-native chitons: broad intertidal habitat range and large densities of Sypharochiton pelliserpentis show no evidence of habitat engineering effect in South Australia

Kiran Liversage* and Jonne Kotta Estonian Marine Institute, University of Tartu, Mäealuse 14, 12618, Tallinn, Estonia Author e-mails: [email protected] (KL), jonne@sea.ee (JK) *Corresponding author

Citation: Liversage K, Kotta J (2019) A rare example of non-native chitons: broad Abstract intertidal habitat range and large densities of Sypharochiton pelliserpentis show no There are numerous examples of non-indigenous rocky-intertidal mobile invertebrate evidence of habitat engineering effect in species, but there have been very few instances of introductions of chitons despite South Australia. Aquatic Invasions 14(2): their often high abundance in fouling assemblages where non-indigenous species 267–279, https://doi.org/10.3391/ai.2019. originate. In 2016, it was observed incidentally that the snake-skin chiton 14.2.07 Sypharochiton pelliserpentis, native to eastern Australia and New Zealand, was Received: 7 August 2018 occurring in Coffin Bay in western South Australia, which is a sheltered bay used Accepted: 4 January 2019 for oyster farming and far outside the previously documented S. pelliserpentis Published: 27 March 2019 range. The main coastal water currents in the region flow from west to east, meaning Handling editor: João Canning-Clode natural larval dispersal into Coffin Bay from eastern Australian populations is Thematic editor: Charles W. Martin unlikely. Surveys of population and community dynamics done in 2018 revealed that a large population has established with the non-native chiton far outnumbering Copyright: © Liversage and Kotta a comparable native chiton (Plaxiphora albida) that may occupy a similar niche This is an open access article distributed under terms of the Creative Commons Attribution License space. The chitons occurred on exposed bedrock habitats as well as cryptic habitats (Attribution 4.0 International - CC BY 4.0). underneath boulders; both species were equally abundant between these habitats OPEN ACCESS. except at one site where S. pelliserpentis densities specifically underneath boulders were 10 times greater than other habitat-types or sites. Sypharochiton pelliserpentis in its native range can largely impact sessile assemblages; here we tested the hypothesis that sessile assemblages would differ on boulders with versus without the non-native chiton, but no evidence of any such effect was found. Sypharochiton pellisperpentis is a common epibiont of oyster-reefs and the origin of its introduction into South Australia may involve transport of oyster-industry materials, which was how introductions of four other non-native benthic invertebrates occurred previously into another Australian region. The geographic extent of S. pelliserpentis in South Australia is unknown at this stage but the chiton was not found during surveys in another nearby bay. Knowledge that oyster-associated chitons may be spreading outside their native ranges, and that they can establish primarily in habitats that are largely hidden from view, such as underneath boulders, can inform monitoring and management practices for ecology of non-indigenous intertidal species.

Key words: intertidal boulder-field, intertidal rocky reef, non-indigenous species, polyplacophora, benthic invertebrate, intertidal grazer, Plaxiphora albida

Introduction Benthic invertebrates are one of the functional groups considered a priority for management of marine non-indigenous species, with numerous incidences of large invasion events. Some benthic marine invasions resulted

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 267 Non-indigenous chiton ecology

Table 1. Review of published literature showing Sypharochiton pelliserpentis densities in its native range in eastern Australia and New Zealand, as well as densities found in the range of South Australia in this study where it is non-native. Densities shown are the means across all contexts (i.e. sites and/or experimental treatments) in each study, and the maximum found from any single context. density*m-2 study region habitat mean maximum New Zealand Ford and Pawley (2008) Waitemata Harbour, Auckland rock shore 27.5 64.0 Coates (1998) Otago Peninsula, Dunedin rock shore 19.0 65.5 Seaward (2006) Kaikoura Peninsula (South Island) rock shore 0.4 0.6 Smith (2009) Kairakau (North Island) rock shore 4.2 5.0 Parker (1976) Great Mercury Island rock shore 113 176 Creese (1988) Leigh Marine Reserve (North Island) rock shore - 19-50 Boyle (1970) Waitemata Harbour, Auckland rock shore 22 50 Australia Blockley and Chapman (2008) Sydney Harbour seawalls 55.2 111.2 Blockley (2007) Sydney Harbour seawalls 28.4 152.8 Blockley and Chapman (2006) Sydney Harbour seawalls 30.0 56.4 Moreira (2006) Sydney Harbour seawalls 72.0 208.4 mean from native range 33.8 84.0 this study Coffin Bay, South Australia rock shore 6.7 28.97

in extensive on-going ecological and economic impacts (Çinar et al. 2014); only very occasionally have they been eradicated (Ferguson 1999; Myers et al. 2000). Important mobile invertebrate invasive taxa have included crabs (Darling et al. 2008; Falk-Petersen et al. 2011; Gribben et al. 2015; Kotta et al. 2018), gastropods (Martel et al. 2004), bivalves (Greene et al. 2011; Möller and Kotta 2017) and sea stars (Byrne et al. 1997). Among marine molluscs, few instances of non-indigenous chitons have been documented. It is uncertain why chitons have rarely become invasive, since they often reach large abundances (Moreira et al. 2007; Nydam and Stachowicz 2007) (Table 1) within the types of fouling assemblages that are prone to exporting non-indigenous species (Baird 1955; Carlton 1999; Darbyson et al. 2009; Frey et al. 2014) and many chitons have widespread dispersal of eggs or larvae (Buckland-Nicks 1993; Yearsley and Sigwart 2011). Adults are often habitat specialists (Kangas and Shepherd 1984; Nikula et al. 2011; Yearsley and Sigwart 2011), however, and these will not likely become established in new regions by relying on anthropogenic disturbances like many other invasive taxa (Bando 2006; Altman and Whitlatch 2007). Adult chitons can be dispersed on artificial structures (Baird 1955; Carlton 1999) and often attach to materials that can become naturally rafted (Nikula et al. 2011). For example, following the 2011 Tōhoku tsunami, six chiton species native to Japan had rafted to Hawaii or the North American coast on tsunami debris (Eernisse et al. 2018). Other examples of non- indigenous chitons of which we are aware include a potentially non-native species on a remote southern Atlantic island (Schwabe and Tsiamis 2017), and Chiton glaucas which is native to New Zealand and is believed to have been introduced to Tasmania (south-eastern Australia) around 1920 (Hayward 1997). This introduction event also included three other invertebrates from New Zealand and is considered to have occurred during transport of live oysters (Clements et al. 2000). A chiton from continental

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 268 Non-indigenous chiton ecology

Figure 1. Photographs of Sypharochiton pelliserpentis observed in Coffin Bay during (a) January 2016 (underneath a boulder that has been overturned), and (b) March 2018 (within shaded depressions on exposed bedrock). The bar at bottom left indicates approximately 1cm. Photos by Kiran Liversage.

Europe and South America was also considered to have been introduced into England in the mid 20th century (Baird 1955). This study quantifies another incident of a chiton apparently becoming established outside its range in southern Australia, which does not appear to have been previously described. The species is Sypharochiton pelliserpentis (Quoy and Gaimard, 1835) which has a documented native range throughout New Zealand, the temperate eastern Australian coast, and Tasmania (Veale and Lavery 2011). To our knowledge, it has never been previously known to occur in any part of the state of South Australia. For example, it is not present in the South Australian marine invertebrate guides of Shepherd and Thomas (1982) or Gowlett-Holmes (2008). Also, we could find no mention of this species from any published rocky intertidal survey in South Australia (e.g. Benkendorff 2005; Dutton and Benkendorff 2008; Janetzki et al. 2015, 2018) including surveys done across large areas specifically targeting chitons (Liversage and Benkendorff 2013). In January 2016, however, a single specimen (Figure 1a) was observed during a scientific survey of fauna on small intertidal boulders in Coffin Bay, western South Australia (K. Liversage, unpublished data) although large populations may have been present at the time on other unsurveyed habitat types. The length of coastline between Coffin Bay and the nearest extent of the chiton’s previously documented native range (Veale and Lavery 2011) is at least 1800 km. Coffin Bay is located west of the previously documented native range, which is the opposite direction of all predominant coastal water currents in the region, which flow eastward (James et al. 2001; Domingues et al. 2007), so a range expansion via dispersal of larvae or natural rafting from its populations in eastern Australia across this distance into Coffin Bay is unlikely. Thus, in the proceeding sections the species is referred to as “non-native” to Coffin Bay. Sypharochiton pelliserpentis can reach large population densities in its native range on natural hard-substrata as well as artificial substrata (Table 1).

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 269 Non-indigenous chiton ecology

Although in Coffin Bay it was first observed underneath an intertidal boulder (Figure 1a), in its native range it is most commonly reported as an exposed-rock species on rock-platform or seawalls (Table 1) but is sometimes found underneath boulders (Liversage 2018). It is considered a strongly interacting species that can change the structure of surrounding sessile assemblages, especially from grazing activity (Luckens 1974; Creese 1988). Although interactions of S. pelliserpentis with other mobile species are currently unstudied, it is likely some form of competition regularly occurs similarly to other chitons that reach such large densities (Duggins and Dethier 1985; Scheibling 1994). On the South Australian coast, there is potential for competition with the native chiton Plaxiphora albida (Blainville, 1825) which occurs in the same habitats and appears ecologically similar (i.e. there may be a large niche overlap). The present study aimed to determine the general extent of establishment of S. pelliserpentis around the town of Coffin Bay, including density measurements of this non-native chiton and of the native P. albida. The survey assessed boulder and bedrock habitat to ensure no habitat-type was unrepresented where the chiton may be occurring. This distinction was incorporated into the hypotheses by testing whether densities of the native and non-native chitons differ on boulders versus bedrock. Lastly, potential effects of this strongly-interacting non-native chiton were investigated by testing the hypothesis that sessile species assemblages will differ on boulders with versus without S. pelliserpentis (i.e. testing whether patterns are evident of impacts at the boulder-scale). Surveys were also replicated in two sites of Boston Harbour, near Port Lincoln, to determine if S. pelliserpentis also occurs now in this nearby bay.

Materials and methods Surveys were done during low tides in March 2018. Four sites were surveyed along the coastline directly adjacent to Coffin Bay township (Figure 2), along with two sites near Port Lincoln, Boston Bay (34°45′05.1″S; 135°51′02.2″E and 34°41′49.5″S; 135°51′18.4″E). Thirty boulders were sampled at each site and were chosen haphazardly for sampling (Chapman 2002), i.e. by walking from 1–10 steps in alternating directions after each boulder was sampled and overturning the closest boulder for the next sample. Care was taken that the same boulder was not sampled more than once. The boulders approximated flattened ellipsoids in shape and were of rough- textured limestone, being interspersed across the boulder fields within a sandy matrix. Numbers of S. pelliserpentis and P. albida were counted and photographs taken of boulder undersurfaces and ruler for scale using an Olympus® TG-4 camera. Densities were later determined by importing photographs into the programme SketchUp v8 (https://www.sketchup.com/) in which 2-dimensional boulder surface areas were measured for density

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 270 Non-indigenous chiton ecology

Figure 2. Map of sampled sites in Coffin Bay, with relative abundances shown of Sypharochiton pelliserpentis (non-native chiton) and Plaxiphora albida (native chiton).

calculations. Bedrock was sampled on consolidated rock surfaces co-occurring among the boulders. Quadrats were of similar size to the average sized boulder (quadrat area = 625 cm2) in which all chitons were counted. Boulders and bedrock were sampled in the mid-low tidal zone of the shore. Sessile assemblages on boulders were measured by randomly superimposing 20 points on each photograph, with the encrusting species under each point representing 5% cover. Many tubeworms occurred on these boulders, but the methods did not allow for differentiation between living and dead tubes, so tubeworms were excluded from the assemblage analysis to prevent counts of dead tubes confounding the analysis on living organisms. A separate analysis was done on only tubeworms (spirorbids and serpulids pooled) which included living and dead animals together, effectively testing for effects on overall structure produced by these tubes, regardless of the status as living or dead. Similarly, the barnacle Amphibalanus amphitrite only occurred as empty tests under boulders so these were analysed separately. The sessile species comparisons were done for boulders but not bedrock, and only at Old Oyster Town and Long Beach, because in the other contexts n ≥ 3 was not able to be reached for

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 271 Non-indigenous chiton ecology

-2 40 Boulders *m

* 30 Bedrock density 20

10

0

S. pelliserpentis 0 Long Beach Snapper Point Old Oyster Town Kellidie Bay Figure 3. Mean (±S.E.) densities of Sypharochiton pelliserpentis occurring underneath intertidal boulders or on nearby exposed bedrock at the four sampled sites in Coffin Bay; n = 30; *Pairwise test P < 0.05.

treatments requiring replicates that have S. pelliserpentis on bedrock (where the species was less common). At Long Beach there were only 5 boulders with the chiton present, so n = 5 was made for all treatments with more available samples by random sample selection, allowing a balanced design (Underwood 1997). Densities of each chiton species were separately compared among sites and between habitat-types using univariate PERMANOVA (based on Euclidian distance) done in PRIMER v7. Homogeneity of variances was tested with PERMDISP, using deviations from the median (equivalent to Levene’s test for univariate analyses; Anderson et al. 2008) and differences among levels of factors were tested with pairwise tests. Multivariate sessile species assemblages were compared between boulders with versus without S. pelliserpentis using PERMANOVA based on Bray Curtis distances. The PERMANOVA routines used 9999 permutations. The separate analyses for tubeworms and barnacles were done using univariate analyses as detailed above. All sessile data were percentage cover so were arcsine transformed prior to analysis (Underwood 1997). “Site” was a random factor in all analyses and other factors were fixed. If any interactions had P > 0.25, they were eliminated to increase the power of the test of the relevant null hypothesis (Underwood 1997).

Results Sypharochiton pelliserpentis was found at all sampled sites around Coffin Bay except one site that was furthest in the bay (Figure 2; this site was excluded from analyses). The native chiton Plaxiphora albida had the same distribution but was always less abundant than the non-native chiton (Figure 2). The greatest S. pelliserpentis densities occurred at the site “Old Oyster Town” where a mean of 29 individuals*m-2 occurred on boulders (Figure 3). This density is approximately one-third the mean maximum density reported from studies across the chiton’s native range, or one-half if only natural rock shores are considered (Table 1). Native P. albida had mean densities ranging from 0.16–2.45 individuals*m-2 across the occupied sites,

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 272 Non-indigenous chiton ecology

Table 2. Univariate PERMANOVA on densities of chitons that are non-native (Sypharochiton pelliserpentis) and native (Plaxiphora albida) in two habitat-types (on exposed bedrock or underneath boulders) and within three sites (random factor) in Coffin Bay (both species were absent from a forth sampled site). Homogeneity of variances was tested using PERMDISP. * P < 0.05, ** P < 0.01, *** P < 0.001. Non-native chiton density Native chiton density df MS F MS F Site = Si 2 2909.44 5.92 ** 90.73 3.56 * Habitat-type = Ha 1 3990.61 0.98 20.66 0.37 Si × Ha 2 4077.36 8.30 *** 55.27 2.17 Res 174 491.10 25.52 PERMDISP P < 0.01 PERMDISP P < 0.01

Table 3. PERMANOVA on multivariate assemblages of living sessile assemblages, and univariate abundances of worm tubes (for which living and dead individuals could not be differentiated) and Amphibalanus amphitrite barnacles (which were all dead tests). The analyses compared between boulders with versus without the presence of the non-native chiton Sypharochiton pelliserpentis, and among three random sites. When the interaction term had P > 0.25 it was eliminated to increase the power of the hypothesis test (Underwood 1997) and is denoted by “–“, but there were no significant terms even with elimination. Homogeneity of variances was tested using PERMDISP. Serpulimorph tubeworms Amphibalanus Living sessile assemblage df (alive + dead) amphitrite tests (dead) MS F MS F MS F Site = Si 1 6949.4 2.27 201.47 3.03 4.71 0.05 Non-native chiton presence/absence = No 1 732.66 0.24 226.71 2.09 289.80 3.11 Si × No 1 2871.4 – 108.64 1.64 4.71 – Res 16 3119.5 66.42 93.27 PERMDISP P > 0.25 PERMDISP P > 0.75 PERMDISP P > 0.5

and was sparsely interspersed among replicates (this prevented attempts at correlation analysis with the S. pelliserpentis). The native and non-native chitons occurred in two habitat-types – underneath boulders (Figure 1a) and on exposed bedrock (Figure 1b). Individuals on exposed rocks sometimes had valves/shells that were heavily overgrown with foliose algae and/or eroded, while those underneath boulders had valves that were more intact although often heavily fouled with non-geniculate coralline algae and spirorbids. Habitat-type effects were interactive with site effects for S. pelliserpentis (Table 2) caused by much greater densities on boulders at one site but no difference at other sites (Figure 2). For the native chiton the only significant effect was from the Site factor (Table 2). In both cases transformations were unable to prevent significantly heterogeneous variances (PERMDISP test), but PERMANOVA was still done as this routine is not sensitive to the assumptions of normality and homogeneous variances when using balanced designs (Anderson and Walsh 2013). On bedrock the mean (S.E.) lengths of P. pelliserpentis and P. albida were 3.28 (0.17) and 4.17 cm (0.49), respectively, and on boulders were 3.24 (0.17) and 3.01 cm (0.68). The most common sessile taxon underneath the boulders was tubes of serpulimorph polychaetes (25.5% cover), followed by nongeniculate coralline algae (7.25%) and Ulva spp. (2.5%). There was no evidence that the sessile assemblages differed between boulders with versus without S. pelliserpentis (Table 3). Lastly, all boulders and quadrats at the two sites surveyed in nearby Boston Harbour did not include S. pelliserpentis or P. albida.

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 273 Non-indigenous chiton ecology

Discussion Transport of live seafood species is an important mechanism that allow non-indigenous marine species to become established in new regions. It is considered likely that this method caused a non-indigenous chiton from New Zealand to become established in Tasmania, with live oysters as the transport medium (Hayward 1997). In the present study it was found that another chiton, Sypharochiton pelliserpentis, had become established in large densities outside its previously documented native range. Taking into account the predominate water currents in the region, it is likely that this chiton can be included among the few known examples of non-indigenous chitons spread through human activities. Coffin Bay is west of the previously known S. pelliserpentis range, while the predominant water currents flow eastward (James et al. 2001; Domingues et al. 2007), which suggests its presence in Coffin Bay may be a human-facilitated introduction and not related to natural dispersal (e.g. from larvae). It is unclear at this stage whether the chiton may be able to spread widely and interfere with important ecological functions or is rather a mostly un- impacting non-native species, and the potential should be considered for a lag phase before any impacts become apparent (Crooks 2005; Simberloff et al. 2013). It has become much more abundant than a similar native chiton; it is unknown, however, if the native species’ relatively low densities are or are not the same as before the new chiton was introduced. There was no evidence that S. pelliserpentis shaped co-occurring sessile assemblages, but this result was preliminary and more in-depth experiments on grazing and “bulldozing” (i.e. unintentional removal of small invertebrates during movement; Safriel et al. 1994) are needed. Sypharochiton pelliserpentis is capable of removing numerous types of algae (Creese 1988; Trowbridge 1995) and barnacles (Luckens 1974), and bioerosion during its grazing has been estimated to contribute 1.7–5.5% of overall coastal rock erosion in some places (Horn 1983), highlighting the potential for large impacts if populations reach high abundances. Analysis of the habitat associations of the native and non-native chitons showed no significant differences in densities of either species between bedrock and boulders, except at the site with greatest S. pelliserpentis densities. The large densities of the non-native species specifically on boulders at this site shows 1) the large amounts of population variability associated with small-scale habitat-type/spatial differences, and 2) the importance of boulder habitat in particular for facilitating this species’ establishment. The relatively large abundances in one habitat of one site may possibly indicate this is the important source population from which larvae are spreading. Most of the patches of bedrock and boulders sampled in this study were in the low shore zone. Many other studies (e.g. those referred in Table 1) have generally documented populations occurring in the mid/high shore zones,

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 274 Non-indigenous chiton ecology

so when occurring as a non-native species in Coffin Bay, a different habitat to usual may be used by this species. Coffin Bay has been a centre for the Australian oyster industry since the mid 19th century, with farming of Pacific oysters Magallana (formerly Crassostrea) gigas beginning in the late 1960’s. S. pelliserpentis is often associated with fouling assemblages on oysters (i.e. occurring in interstices among the oyster shells; Bugnot et al. 2015) and it could be considered likely that the chiton was introduced to Coffin Bay via transport of oysters, similarly to how Chiton glaucus was introduced to Tasmania (Hayward 1997). The chiton does not appear to be included in any previous species lists of the bay (e.g. the Coffin Bay species identification guide; Saunders 2009) raising the possibility that the introduction has been recent, in which case any deleterious ecological effects caused by the chiton may not yet have appeared. Delays or lags in impacts are a common feature of species invasions (Crooks 2005; Simberloff et al. 2013); the fact that S. pelliserpentis is known to have strong ecological impacts in other regions, and appears to have only recently established in Coffin Bay, raises the possibility of such delayed ecological impacts. In addition, we assessed only potential impacts on co-occurring sessile species. A range of other impacts may also occur, for example, foliose algae on exposed hard surfaces (which provides habitat for many native species; e.g. Thomsen et al. 2016) may be impacted. Competition among rocky intertidal grazers is often extreme (e.g. Underwood 1984) and changing competition dynamics with native chitons, limpets and gastropods in Coffin Bay could also be assessed in further research. The overall geographic extent of spread of S. pelliserpentis in South Australia is currently unknown. It may extend further toward the entrance of Coffin Bay, and almost certainly will occur on the nearby northern bay side. The complete absence of S. pelliserpentis or P. albida in the site furthest in the bay (Kellidie Bay; Figure 2) may indicate highly sheltered inner bay areas are unsuitable habitat for these species. The shallow and sheltered conditions in this bay may lead to high temperatures and salinities during summer months with potential consequences such as reduced species richness. Sypharochiton pelliserpentis was not observed at two sites surveyed in nearby Boston Harbour, but further surveys should test if it occurs in South Australian bays besides Coffin Bay. In conclusion, it is confirmed that a large population of a non-native chiton has become established in Coffin Bay. The species has strong ecological interactions in its native range, and its presence in Coffin Bay may result in changes to ecological systems there, although so far there is no evidence of this based on our present results. There has been increasing interest in transport of non-native chitons across regions following discovery that a range of species can be transported during episodic rafting opportunities (Eernisse et al. 2018), although the dynamics of potential

Liversage and Kotta (2019), Aquatic Invasions 14(2): 267–279, https://doi.org/10.3391/ai.2019.14.2.07 275 Non-indigenous chiton ecology

establishment in recipient ecological systems are currently unknown. This study provides an example of a non-native chiton, potentially transported via oyster farming activity, successfully establishing in a new ecological system. More research on its habitat requirements may reveal the causes of large variability associated with different sites and habitat-types. Other research opportunities could include genetic analysis to determine where the non-native chiton originated from (e.g. from native populations in Tasmania or the neighbouring region, Victoria, or possibly where the previous non-native chiton (Chiton glaucus) originated from in New Zealand). Genetic analysis has been done for native S. pelliserpentis populations in New Zealand (Veale and Lavery 2011) and extending those methods to southern Australia may assist in understanding the ecological dynamics surrounding this non-indigenous chiton and thus gaining information to prevent further spread of non-indigenous marine species that may impact native intertidal communities.

Acknowledgements

Research by K.L. was funded by the European Regional Development Fund (Programme Mobilitas Pluss, project number MOBJD31). The work was supported by the Estonian Research Council (Institutional Funding, IUT02-20). We appreciate the comments and suggestions provided by Karen Gowlett-Holmes, and one other anonymous reviewer.

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