Mauritius kestrel

The Mauritius kestrel (Figure 2.1) is the only surviving resident species of raptor in Mauritius, and the population is thought to have been isolated for 1.9-2.6 million years (Groombridge et al., 2002). It is an Accipiter-like falcon that typically inhabits native forest (Temple, 1987; Cartwright, 2011), but has shown flexibility in occupying agricultural mosaics with a forest component (Burgess, 2008; Cartwright, 2011). It hunts a variety of introduced small birds, mammals and reptiles, and some insects, but shows a preference for native Phelsuma geckos (Jones, 1987; Temple, 1987; Carter and Jones, 1999). It usually forms monogamous pairs, which defend a breeding territory (Jones, 1987), and Mauritius kestrels are able to breed during their first year. Clutches of 1-5 eggs are laid from September onwards and breeding activity continues into the summer months, with the last young

fledging in the middle of the summer rainy season (Nicoll, 2004; Cartwright, 2011). It nests in natural cavities in cliffs or trees. It is primarily the female that incubates the eggs and broods the nestlings, while the male provisions the family until the young have fledged (Cartwright, 2011; Jones et al., 2013a). The species is usually single-brooded, but occasionally second, and rarely third, clutches are laid if the first attempt fails or a brood fledges early enough in the season (Cartwright, 2011). The breeding season spans two calendar years and is referred to using both of those years, e.g. 2014/2015.

It is thought that the species was widely distributed in pristine forested Mauritius before colonisation by humans (Jones, 1987). However, it suffered severe declines owing to habitat destruction, invasive non-native species and pesticide use. Nests and fledglings are predated by invasive mammal species, including black rats (Rattus rattus), small Indian mongooses (Herpestes auropunctatus), feral cats (Felis catus) and crab-eating macaques (Macaca fascicularis) (Cade and Jones, 1993). By 1949, when the implementation of organochlorine pesticides for malaria eradication took place, the species was already restricted to upland areas (Safford and Jones, 1997), but pesticide exposure has been implicated in a further crash in the population (Jones, 1987; Safford and Jones, 1997). By the 1960s, the species was restricted to southwestern Mauritius (Jones, 1987), the only forested area not targeted for pesticide use (Cheke, 1987a). The species experienced a severe population bottleneck and

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neared extinction in the mid-1970s, reaching a low of only four wild individuals, probably including only one breeding pair (Temple, 1986; Jones, 1987; Jones et al., 2013a), with two more individuals in captivity (Cade and Jones, 1993). Subsequent genetic analysis provides supporting evidence of a severe population bottleneck, with a probable low of four birds (Groombridge et al., 2001).

The Mauritius kestrel has since become an icon of species recovery and conservation. A recovery programme was initiated in 1973, which has involved captive breeding, cross-fostering in wild nests, releases of captive-bred and reared birds, clutch manipulation, provision of supplementary food and nest-boxes, and control of predators (Cade and Jones, 1993; Jones et al., 1995). The species was reintroduced to upland forested areas in the Bambous and Moka mountain ranges, and additional birds were released to augment the remnant population in the Black River Gorges (Jones et al., 1995; Nicoll, 2004; Cartwright, 2011). Following its reintroduction in 1991, the population in the Moka Mountains died out in the early 2000s (Jones et al., 2013a).

Figure 2.1. Mauritius kestrel from the Bambous Mountains population.

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Figure 2.2. Mauritius kestrel population trend and releases in the Bambous Mountains, Mauritius, from 1987/1988 to 2014/2015: black points and lines = number of monitored breeding pairs; grey columns = number of birds released. Year indicates that of the start of the breeding season.

The species became extirpated from the Bambous Mountains in the 1950s (Cade and Jones, 1993), but was reintroduced from 1987 onwards, with releases carried out between 1987 and 1994 (Nicoll, 2004) (Figure 2.2). Intensive management of the Bambous Mountains population continued until the 1993/1994 season (Nicoll, 2004). The re-established population has since stabilised at ~50 breeding pairs (Figure 2.2). While the Bambous Mountains population is stable, the smaller extant population in the Black River Gorges has shown recent declines (BirdLife International, 2018a).

The two extant populations rely heavily on the continued provision and maintenance of artificial nest-boxes, especially in the Bambous Mountains, where there is a shortage of suitable natural cavities (Cartwright, 2011). The species shows very limited dispersal, with most birds staying within 2 km of their natal territory (Nevoux et al., 2013). Natural

recolonisation may be limited by conspecific attraction, in which transient non-breeders tend to occur at the margins of breeding territories, rather than occupying more remote patches of habitat (Jones et al., 1995; Cartwright, 2011). This could mean that further population growth will not be possible without large-scale forest restoration (Cartwright, 2011). The two extant populations are separated by over 18 km and are effectively isolated from one-another (Cartwright, 2011). The species is now estimated to number ~250 individuals (M. A. C. Nicoll,

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pers. comm.). Research has confirmed that the population is inbred and has lost substantial genetic diversity, although the long-term consequences for the species are uncertain (Groombridge et al., 2000; Groombridge et al., 2001; Ewing et al., 2008). The species is currently listed as endangered on the IUCN Red List (BirdLife International, 2018a).

Previous research has shown that the species is sensitive to variability in rainfall, with breeding success and juvenile survival negatively affected by increased rainfall in the rainy season, from December onwards (Nicoll et al., 2003; Nicoll, 2004). Breeding phenology is influenced by rainfall in spring, with delayed egg-laying in response to more frequent rainfall (Nicoll, 2004; Senapathi, 2009; Senapathi et al., 2011). Such delays potentially increase the exposure of broods to heavy rainfall in December, which reduces breeding success (Nicoll, 2004; Senapathi, 2009; Senapathi et al., 2011). However, there is no evidence that population dynamics are affected by rainfall variability (Senapathi, 2009). Research has found spatial heterogeneity in recruitment rates (Burgess et al., 2008), avoidance of agricultural areas by juveniles (Burgess et al., 2009), and reduced egg to fledging survival in territories with a higher agricultural component (Burgess et al., 2011). Increased exposure to agriculture worsens the seasonal decline in breeding fitness and the life-history strategies of birds are altered by exposure to a high agriculture component in their natal territory (Cartwright, 2011; Cartwright et al., 2014a, b). The population has also provided opportunities to study the mechanisms through which density-dependence occurs (Nevoux et al., 2011) and the individual fitness consequences of dispersal (Nevoux et al., 2013).