Data Zone
Justification of Red List category
This species has an extremely small range and population size, restricted to a single subpopulation on the southwestern slope of Mauna Kea where its survival is tightly linked to the availability of m?mane seed pods. Annual transect surveys indicate that the species is declining at a rate exceeding 50% in three generations, driven primarily by prolonged drought and habitat degradation by introduced ungulates and invasive plants. The species is restricted to a single location where all individuals present are highly vulnerable to fire. It is therefore assessed as Critically Endangered.
Population justification
In 2021 the population was estimated at 452?940 individuals (point estimate of 678), with previous estimates of 1,030?1,899 individuals (point estimate of 1,432) in 2019, and 964?1,700 individuals (point estimate of 1,312) in 2020 (Genz et al. 2022). Although 2021 abundance estimates are approximately 50% of 2020 estimates, this is considered biologically unlikely and previous low estimates from annual surveys in 2000, 2010 and 2015 have been immediately followed by an increase in subsequent years (Genz et al. 2022). The five-year average (2017-2021) population size is 1,128 individuals, and based on an approximate estimate that 86% are mature individuals (C. Farmer in litt. 2016) this equates to 970 mature individuals. The population size is therefore tentatively placed in the range 800-1,200 mature individuals.
Trend justification
During 1998?2005, estimates fluctuated between 4,000 and 6,000 individuals, followed by a steep decline - between 2005 and 2008 the population fell from c.5,500 to c.1,800 individuals (Genz et al. 2022). This steep decline levelled off somewhat after 2010, but trend assessments show that the population has continued to decline, leading to a record low in 2021 (Genz et al. 2022). The mean decline during 1998?2021 was 229 individuals per year, resulting in an 89% decline in the population over the entire 23-year monitoring period (Genz et al. 2022), equivalent to c.70% within three generations. Within the most recent three-generation period (12.5 years; Bird et al. 2020), annual surveys indicate that the population declined by approximately 50-60% (Genz et al. 2022) and this is thought to continue given that the threats are ongoing. Although the size of the overall area containing detections of the species has not shown substantial change since 1998 (Genz et al. 2022), there has reportedly been a decline of detections within sections of this area which has continued in the 2022-2023 annual surveys (C. Farmer in litt. 2023).
This species is restricted to the island of Hawaii in the Hawaiian islands (U.S.A.), where it was abundant, although locally distributed, until the beginning of the 20th century; evidence from the fossil record suggests that the species occurred throughout the archipelago prior to human settlement (Olson and James 1982, Burney et al. 2001). In 1997, it occupied an estimated 78 km2 and numbered 4,396 individuals, mostly on the western slope of Mauna Kea, where 20.5 km2 was estimated to hold 72% of the total population (Scott et al. 1986, Fancy et al. 1997, Banko et al. 1998). The species' range remains centred on the western slope, and it has contracted such that c.96% of the population is found in 4,600 ha of forest (R. Camp in litt. 2014). It has not been found in annual surveys on the east slope since 2004 (P. Banko in litt. 2007, C. Farmer in litt. 2007, R. Camp in litt. 2014, 2016). A small colony of reintroduced individuals has been extirpated on the northern slope (Banko et al. 2014a).
It is confined to altitudes of c. 2,000-3,000 m, favouring dry māmane and māmane-naio forest. It feeds primarily on māmane seeds, flowers, and insects (Banko et al. 2002, Banko and Farmer 2014, Hess et al. 2014), with the availability of māmane seeds affecting productivity and adult survival. The species is morphologically and behaviourally adapted to feeding on māmane seeds, grasping pods with its feet and using its stout bill to tear them open. In drought years, most birds do not attempt to breed (Jacobi et al. 1996, Pratt et al. 1997). The species exhibits low rates of reproduction (Banko et al. 2002, Banko and Farmer 2012), laying fewer eggs and taking longer to raise its young compared with mainland songbirds (Hess and Banko 2006).
The most significant declines in this species' range and population are thought to have been caused by human-induced habitat loss and degradation and impacts from feral and domestic ungulates (Banko et al. 2002, Banko and Farmer 2014, Banko et al. 2014b). Māmane forest is slow to recover and young māmane requires about 25 years of growth before the species frequent them for foraging and nesting (P. Banko in litt. 2023).
The subalpine forest habitat of this species has been severely overbrowsed by sheep, mouflon sheep, goats, cattle, and its nests and adults are preyed upon primarily by feral cats Felis catus, but also by introduced Black Rats Rattus rattus, Short-eared Owl Asio flammeus, and rarely Hawaiian Hawk Buteo solitarius (Banko et al. 2002, Banko and Farmer 2014). Up to 11% of nests are depredated by feral cats each year (Hess and Banko 2006). Grazing by cattle was a historical factor in the species' decline, although cattle are now limited to pastures that are unsuitable for L. bailleui (C. Farmer in litt. 2007, Banko et al. 2014b). Continuing threats include grazing by feral sheep, wild sheep Ovis gmelini, and their hybrids, which slows māmane regeneration (Pratt et al. 1997, Banko and Farmer 2012, Banko et al. 2013, Banko et al. 2014b), and alien insects preying on and parasitising native insects (Pratt et al. 1997, Banko and Farmer 2012), particularly at low elevations (Banko and Farmer 2012). Native caterpillars are an important source of protein for nestlings, and possibly breeding females; however, they are preyed upon by Yellow-jacket Wasps and several ant species, particularly Argentine Ants Linepithema humile, whilst parasitoid wasps kill the caterpillars by laying their eggs on or inside them (Banko and Farmer 2012). Damage of Myoporum sandwichense tress from Naio thrips Klambothrips myopori may lead to increased tree death and increased fire risk (USFWS 2020). Alien grass cover such as African Fountain Grass Cenchrus setaceus and Orchard Grass Dactylis glomerata is high in much of the species' range (Banko and Farmer 2014), suppressing māmane regeneration and increasing the threat of fire (Thaxton and Jacobi 2009). Alien shrubs and vines potentially displace māmane and other native plants; e.g., Cape Ivy Delairea odorata and Common Gorse Ulex europaeus (Banko et al. 2020). Root rot caused by the non-native fungus Armillaria mellea is also responsible for dieback occurring in māmane forests (Banko et al. 2014b).
Increasing human activities, such as military training, further increase the chances of fire (Thaxton and Jacobi 2009). In 2006-2007, there were numerous fires on and near Mauna Kea, and fires in August and September 2010 affected c.560 ha of suitable habitat on the southern slope of the mountain (American Bird Conservancy 2010, Banko et al. 2014b). A fire in the species' core area could potentially affect >90% of the population (C. Farmer in litt. 2016). The opening of trails for all-terrain vehicles in the Mauna Kea Forest Reserve is a concern (C. Farmer in litt. 2007), and may cause disturbance and habitat degradation.
In addition to the aforementioned threats, this species' very restricted range means it could be extirpated by extreme events such as drought and storms (Banko and Farmer 2012), and drought has contributed to declines given its high dependence on the unripened māmane seeds (Lindsey et al. 1997, Banko et al. 2013). Drought conditions on Mauna Kea occurred during 74 percent of the months from 2000 to 2010, recorded in all but two months from 2006 to 2010 (Banko et al. 2013). This leads to reduced māmane seed production, increased tree death and increased fire risk (VanderWerf 2012, USFWS 2020). Further declines in breeding success are expected in line with drought induced by El Niño, conditions of which are present as of 2023 and expected to increase (P. Banko in litt. 2023). This species' low reproductive capacity means it is slow to recover from perturbations. Demographic patterns of māmane mortality are under investigation, as māmane may be under threat from pathogens (USFWS 2006, Banko et al. 2014b).
Climate change poses a long-term future threat to the species, as drought frequency and intensity are expected to increase at higher elevations, further affecting vegetation structure and compositions, and reducing the habitat's carrying capacity (Benning et al. 2002, Giambelluca and Luke 2007, Banko et al. 2013). Given that this species occurs primarily at elevations where mosquitoes are absent, it is thought to be more limited by habitat availability than by avian disease (Kilpatrick 2006). However, avian malaria and avian pox have significantly affected other forest birds and are likely to prevent the species re-establishing in areas of its former range (Jacobi et al. 1996). Additionally, climate change is projected to cause an increase in the elevation below which regular transmission of avian malaria occurs, reducing the area of suitable habitat and exacerbating existing declines (USFWS 2020). According to climate projections, due to increased rainfall and temperatures, high elevation areas for Hawaiian bird populations will only remain free of mosquitos to mid-century (Liao et al. 2015) and GIS simulation has shown that a 2 degrees Celsius increase would cause a c.100% decline in the land area where transmission is currently only periodic (Benning et al. 2002). Palila are highly sensitive to the disease and if the transmission zone rises sufficiently it is entirely plausible that the entire species would be affected by malaria (C. Farmer in litt. 2016, 2023).
Conservation and Research Actions Underway
The species' population has been monitored since 1980 (Scott et al. 1986, Leonard et al. 2008, Camp et al. 2016). Annual surveys are conducted to monitor range and abundance (USFWS 2020). In 1979, 1987 and 1998, federal courts ordered the eradication of feral goats and feral and wild sheep species from the species' habitat on Mauna Kea, and these rulings have remained in effect despite six legal challenges (Banko et al. 2009, 2014b). Forest regeneration has improved as a result, although current efforts to reduce sheep have not been sufficient to allow the complete recovery of māmane forests (Banko et al. 2014b). Palila have some site fidelity, but have also been detected moving around Mauna Kea into new forested areas (Banko and Farmer 2014). In 1993, some birds were translocated to a new site on the eastern slope where predators were controlled and, although many homed back to their capture site, at least two pairs stayed and bred successfully (Fancy et al. 1997). By the end of 2006, 188 wild birds had been translocated from the western to the northern slope of Mauna Kea (Banko and Farmer 2014). Approximately 34% persisted for longer than two months, and a small colony of up to c.23 birds remained on the northern slope and bred successfully until 2010. Although F2 generation offspring were observed, the colony was not sustainable without additional management (Banko et al. 2009, Banko and Farmer 2014) and it was extirpated by 2011 (P. Banko in litt. 2012). Egg-laying occurred in 2004, and independent juveniles were produced in every year from 2005-2010 (Banko and Farmer 2014). This translocation programme was aided by a captive breeding programme initiated at the Keauhou Bird Conservation Center in 1996 (USFWS 2006, Banko and Farmer 2014). Of 28 captive-reared birds released in 2003-2009, at least 10 persisted in the reintroduction area for at least one year, with 2 males successfully producing fledglings with translocated wild females in the north slope colony before its extirpation (Banko and Farmer 2014). Results suggest that, although some translocated birds nest successfully and others remain in the new area for over a year, eventually all birds return to the source area (western slope) (Banko et al. 2020). Eight captive birds were released at the Puʻu Mali Restoration Area on the north slope of Mauna Kea in 2019 (USFWS 2020). After being temporarily held in release aviaries, individuals dispersed widely but most died from predation or disappeared within one week (USFWS 2020). A captive population is maintained by San Diego Zoo Global (USFWS 2020).
The construction of a highway through unoccupied, federally designated critical habitat was approved in 1999. A mitigation plan accompanied the development, including the suspension of cattle grazing in pastures adjacent to the species' range on the northern and western slopes. The species' conservation was the subject of detailed research, and funding from the mitigation plan supported translocation research and enabled the expansion or continuation of studies into the species' ecology and limiting factors, māmane ecology, food availability, predator ecology and management, and fire ecology (Banko and Farmer 2014). The state of Hawaii continues habitat restoration and research into restoration methods (Banko and Farmer 2014). Work is being carried out to restore habitat by controlling African Fountain Grass Pennisetum setaceum, which increases the frequency and intensity of fires, and Cape Ivy Delairea odorata, which reduces the vigour of native trees (American Bird Conservancy 2010). In 2010, a comprehensive fire management plan was being developed for the Mauna Kea area (Thaxton and Jacobi 2009, American Bird Conservancy 2010). There are currently 100 km of fencing around critical habitat and around the Kaʻohe and Puʻu Mali restoration areas, as well as 26 km of fire breaks in place (USFWS 2020). The Mauna Kea Forest Restoration Project conducts weed control, forest restoration, forest monitoring, fence monitoring and maintenance, community outreach, volunteer trips, and ungulate control at Kaʻohe and Puʻu Mali Restoration Areas and in the core habitat within Mauna Kea Forest Reserve, and predator control is ongoing (USFWS 2020). Research is ongoing to determine survival rates, causes of nest failure, habitat use, dispersal, nest site characteristics, disease prevalence and predator prevalence (USFWS 2020).
19 cm. Large finch with short, rounded bill. Male has golden-yellow head and breast surrounding black lores and bill, dark grey back and rump, white underparts, and dark wing and tail feathers with broad, golden edges. Female less golden and with grey of back extending forward on hindneck to crown. Similar spp. Introduced yellow morph of House Finch Carpodacus mexicanus has streaks on back and belly, yellow on rump. Introduced Yellow-fronted Canary Serinus mozambicus has grey nape, yellow underparts and rump, and bold facial markings. Both smaller. Voice Quiet, sweet canary-like song. Call a sweet chee-klee-o or pa-lee-la. Hints Best looked for at Pu'u La'au on Big Island.
Text account compilers
Vine, J.
Contributors
Baker, H.C., Baker, P.E., Banko, P., Camp, R., Farmer, C. & Pratt, T.
Recommended citation
BirdLife International (2024) Species factsheet: Palila Loxioides bailleui. Downloaded from
https://datazone.birdlife.org/species/factsheet/palila-loxioides-bailleui on 22/11/2024.
Recommended citation for factsheets for more than one species: BirdLife International (2024) IUCN Red List for birds. Downloaded from
https://datazone.birdlife.org/species/search on 22/11/2024.