CPS49 in chicken embryos causes loss of vessels before changes in limb signaling pathways. In zebrafish embryos, CPS49 causes rapid loss of vessels, which from HUVEC studies is due to inhibition of actin cytoskeletal dynamics (Therapontos et al., 2009). Other studies in zebrafish have demonstrated that thalidomide alters key molecules in vascular development, including vascular endothelial growth factor VEGF, a key signaling molecule in vessel development (Yabu et al., 2005). Another study in zebrafish indicates thalidomide may affect gene expression patterns of Shh and Fgf8 before vessels form in developing fins (Ito et al., 2010). However vascular markers, including VEGF, were not looked at in this study, and vessels in other regions of the embryo were not detailed. As zebrafish fins and chicken limbs develop and form differently, this could reaffirm that the drug behaves differently in different species.
Finally, many important molecules involved in embryonic vascular development and patterning have been reported to exhibit changed expression patterns following thalidomide exposure, for example, actin and tubulin, integrins, vascular endothelial growth factor, PDGFβ, nitric oxide, ceramide, angiopoietins, Notch, HIF, Slit2/Robo signalling and ROS (Stephens et al., 2000; Vargesson, 2009, 2013; Feng et al., 2014; Li et al., 2014.)
accidental growth mika tan
Framework of thalidomide induced embryonic damage. This framework incorporates the majority of the previously proposed models/hypotheses to attempt to provide an explanation for thalidomide embryopathy. Thalidomide and/or a breakdown product after binding a molecular target acts negatively on smooth muscle negative blood vessels, likely affecting the actin cytoskeleton of the endothelial cells, and preventing their proliferation and migration into avascular regions, causing oxidative stress, cell death, and gene expression loss, resulting in tissue damage. In rapidly developing tissues and organs, such as the limbs and internal organs, this would be devastating, causing tissue loss or tissue function loss, preventing growth. The damaged or missing tissues would then also fail to properly recruit and pattern proper chrondrogenesis, nerve innervation, muscle patterning, etc., exacerbating the condition and damage.
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.
In addition to Euglandina spp. and Platydemus manokwari, a number of other species have been deliberately introduced outside their native ranges as biological control agents, including the snails Edentulina spp., Gonaxis kibweziensis and G. quadrilateralis (e.g. Cowie 1997, 2000, 2001a). There are also numerous other non-native invertebrate species that prey on or parasitise land snails and that have been introduced outside their native ranges, most commonly accidentally through human activities. These include predatory snails such as Gulella bicolor, Streptostele musaecola and Oxychilus alliarius (Cowie 1997; Meyer and Cowie 2010; Curry et al. 2016, 2020), various flatworms, carabid beetles, ants, phorid and sarcophagid flies, nemertean ribbon-worms, and not inconsiderable numbers of helminth parasites (Barker 2004). For the most part, the impact of these non-native species on land snail communities has yet to be examined in any depth and is beyond the scope of this review.
Lissachatina fulica was established on the Maldives by 1957 when it was present on Addu atoll (Mead 1961). Euglandina have not been introduced but Platydemus manokwari was deliberately introduced to six islands, including Favahmulah (Fua Mulaku) and the atolls of Addu and Male in 1985 (Muniappan 1987) (also see Eldredge and Smith 1994, 1995). On Favahmulah P. manokwari spread up to a radius of 180 m within approximately! year and was claimed to have controlled the L. fulica population, even though Muniappan (1987) acknowledged that snail declines had occurred on islands where flatworms appeared not to have become established. The flatworm was photographed on Rasdhoo Atoll in 2017 (A. Pichler pers. comm.), possibly following accidental transport between the islands; L. fulica was highly abundant on Rasdhoo at the time.
Platydemus manokwari had spread to urban Manila by 1985, apparently through accidental introductions. It was still present in 2019 and L. fulica remained abundant in the Manila region (Constantino-Santos et al. 2014), refuting the claim by Waterhouse and Norris (1987) that it had controlled the L. fulica population. There are subsequent records from Mindanao and Cebu in 2020 (Muico 2020).
Platydemus manokwari has been recorded from Queensland (1976) and the Northern Territory (2002) (Winsor 1990; Justine et al. 2014). There have been accidental movements of P. manokwari by people (e.g. in north Queensland, from Bluewater to Weipa, Cape York; Waterhouse and Norris 1987) and these probably account for the additional records of the species in Bowden and Brisbane in Queensland. It is restricted to urban gardens and has not spread into drier natural habitats, except at Lake Eacham, Atherton Tablelands, where gardens abut rainforest. This invasive population seems constrained by surrounding dry farmland (Winsor pers. obs.).
However, it was other species of snails introduced to Bermuda that were targeted by a biocontrol programme. These were two Mediterranean species, primarily Otala lactea, imported for food in 1928, and Rumina decollata, introduced accidentally with plants in the late 1870s (Bennett and Hughes 1959; Simmonds and Hughes 1963; Bieler and Slapcinsky 2000). Based on the supposedly successful but ultimately ill-conceived, unsuccessful and catastrophic attempts to control L. fulica with predatory snails in Hawaii, the Commonwealth Institute of Biological Control deliberately released Euglandina (and other predatory species) in Bermuda in 1958 and 1960 to control O. lactea and R. decollata, which by then were considered agricultural pests (Simmonds and Hughes 1963). Euglandina has since spread throughout much of the larger islands of Bermuda (Bieler and Slapcinsky 2000). Although control of O. lactea was reported in 1962 by the Bermuda Department of Agriculture, ultimately control of neither species was achieved (Bieler and Slapcinsky 2000). However, Euglandina may have been responsible for the decline of endemic species, notably species of the endemic genus Poecilozonites, which began in the 1960s (Bieler and Slapcinsky 2000). By 2004 the two non-fossil Poecilozonites species were thought to be on the verge of extinction and individuals were brought into captivity to establish a captive rearing programme, which has become highly successful to the extent that large numbers of P. bermudensis have already been released back into the wild on the offshore Nonsuch Island, a reserve free of Euglandina (Outerbridge et al. 2019).
The spread of Euglandina species is thought to be entirely due to deliberate introduction, mostly official introductions by government agencies, but also unofficial transport by private individuals within countries. No official introductions have been recorded since 1993, although some transport within island groups has occurred since then. Introduction of P. manokwari may originally have been accidental; its probable movement from New Guinea to Australia in the 1970s and from there to Guam by 1978 predated the suggestion that it could be used as a biocontrol agent. Most introductions of P. manokwari appear to have been accidental, probably through the transport of plant material. Once deliberate releases occurred, multiple populations became potential introduction sources and the rate of documentation of new invasions is increasing (from two introductions in the 1970s to 17 since 2010), although this may in part be the result of increased publicity. As P. manokwari has only been identified as a species of concern since 1992 (Hopper and Smith 1992) and the focus of attention recently, it is likely to have been under-recorded and may have a much wider distribution than is currently known. Detection of P. manokwari may be relatively more difficult than of Euglandina as it does not leave behind shells when it dies.
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