Abstract
Since 1955 snails of the Euglandina rosea species complex and Platydemus manokwari flatworms were widely introduced in attempted biological control of giant African snails (Lissachatina fulica) but have been implicated in the mass extinction of Pacific island snails. We review the histories of the 60 introductions and their impacts on L. fulica and native snails. Since 1993 there have been unofficial releases of Euglandina within island groups. Only three official P. manokwari releases took place, but new populations are being recorded at an increasing rate, probably because of accidental introduction. Claims that these predators controlled L. fulica cannot be substantiated; in some cases pest snail declines coincided with predator arrival but concomitant declines occurred elsewhere in the absence of the predator and the declines in some cases were only temporary. In the Hawaiian Islands, although there had been some earlier declines of native snails, the Euglandina impacts on native snails are clear with rapid decline of many endemic Hawaiian Achatinellinae following predator arrival. In the Society Islands, Partulidae tree snail populations remained stable until Euglandina introduction, when declines were extremely rapid with an exact correspondence between predator arrival and tree snail decline. Platydemus manokwari invasion coincides with native snail declines on some islands, notably the Ogasawara Islands of Japan, and its invasion of Florida has led to mass mortality of Liguus spp. tree snails. We conclude that Euglandina and P. manokwari are not effective biocontrol agents, but do have major negative effects on native snail faunas. These predatory snails and flatworms are generalist predators and as such are not suitable for biological control.
Highlights
Species extinction rates are currently higher than background levels, representing a sixth mass extinction event (Waters et al 2014; Ceballos et al 2015)
This high extinction in land snails may be representative of invertebrates as a whole (Regnier et al 2015a), but this is difficult to ascertain because of the lack of data for the vast majority of invertebrate species (Bouchet et al 2002; Brodie et al 2017)
Lissachatina fulica was established in the Comoros in the nineteenth century (Mead 1961) or earlier (Bequaert 1950)
Summary
Species extinction rates are currently higher than background levels, representing a sixth mass extinction event (Waters et al 2014; Ceballos et al 2015). The levels of extinction are not evenly spread across taxa with high levels reported for land snails (Lydeard et al 2004; Regnier et al 2009, 2015a, b) estimated overall at 10% loss and an extraordinary 75% extinction of Polynesian island land snails (Regnier et al 2015a) This high extinction rate is probably due in part to many snail species having very restricted geographical ranges, combined with low mobility, making them very sensitive to changes in land use, habitat quality and/or a range of non-native predators (Lydeard et al 2004; Regnier et al 2015b; Chiba and Cowie 2016; Cowie et al 2017). This is despite growing international recognition that our ability to partition species losses among multiple drivers is critical to advancing our understanding and mitigation
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